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373
Chapter 7
Heavy Metals
Contents
7.6.3. Freshwater ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . 411
7.6.3.1. Metals in freshwater . . . . . . . . . . . . . . . . . . . . . . 411
7.0. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 374
7.6.3.2. Metals in freshwater sediment . . . . . . . . . . . . . . . 412
7.6.3.2.1. River and lake bottom sediments . . . . . 412
7.1. Physical/chemical characteristics . . . . . . . . . . . . . . . . . . . . . 374
7.6.3.2.2. Freshwater particulates . . . . . . . . . . . . 414
7.1.1. Identification of metals to be considered . . . . . . . . . . . . . . 374
7.6.3.2.3. River heavy metal fluxes . . . . . . . . . . . 414
7.1.2. Speciation of metals in the environment . . . . . . . . . . . . . . 375
7.6.3.3. Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . 415
Atmosphere . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 375
7.6.3.4. Algae and plants . . . . . . . . . . . . . . . . . . . . . . . . . 415
Aquatic systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 375
7.6.3.5. Metals in freshwater invertebrates . . . . . . . . . . . . 415
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 376
7.6.3.6. Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 415
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 376
7.6.3.7. Metals in aquatic birds . . . . . . . . . . . . . . . . . . . . 416
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 376
7.6.3.8. Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 416
7.2. Sources of pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 376
7.6.4 Wetland ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 416
7.2.1. Natural sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 376
7.6.5. Marine ecosystem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 416
7.2.2. Anthropogenic sources . . . . . . . . . . . . . . . . . . . . . . . . . . . 377
7.6.5.1. Seawater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 417
7.2.2.1. Sources and fluxes of atmospheric input
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 417
on a global scale. . . . . . . . . . . . . . . . . . . . . . . . . . 377
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 417
7.2.2.2. Sources and fluxes of aquatic input on a global scale 379
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 417
7.2.2.3. Terrestrial input and output of heavy metals
7.6.5.2. Sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 418
on a global scale . . . . . . . . . . . . . . . . . . . . . . . . . 380
Copper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 420
7.2.3. Emission inventories for sources within and outside
Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 420
the Arctic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 380
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 420
7.2.3.1. Atmospheric emissions from sources outside
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 420
the Arctic and their trends . . . . . . . . . . . . . . . . . . 381
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 420
7.2.3.2. Atmospheric emissions from sources within
Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 421
the Arctic. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 382
7.6.5.3. Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . 421
7.2.3.3. Aquatic emissions from sources outside the Arctic 383
7.6.5.4. Algae . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 421
7.2.3.4. Aquatic emissions from sources within the Arctic . 384
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 421
7.2.3.5. International agreements on emission reduction . 384
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 421
Mercury and selenium . . . . . . . . . . . . . . . . . . 422
7.3. Special issues of pathways of metals . . . . . . . . . . . . . . . . . . 384
7.6.5.5. Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . 422
7.3.1. Atmospheric transport . . . . . . . . . . . . . . . . . . . . . . . . . . . 384
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 422
7.3.1.1. Dispersion models to study the impacts of
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 423
sources outside the Arctic. . . . . . . . . . . . . . . . . . . 384
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 423
7.3.1.2. Receptor models to study the impact of sources
Selenium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 424
outside the Arctic . . . . . . . . . . . . . . . . . . . . . . . . . 386
7.6.5.6. Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 424
7.3.1.3. Modeling the dispersion of emissions from
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 424
sources in the Arctic . . . . . . . . . . . . . . . . . . . . . . . 387
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 424
7.3.2. Rivers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 387
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 425
7.3.3. Estuaries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 387
Selenium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 426
7.3.4. Oceans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 388
7.6.5.7. Seabirds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 426
7.3.5. Ice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 388
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 426
7.4. Toxicological characteristics . . . . . . . . . . . . . . . . . . . . . . . . . 388
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 426
7.4.1. Toxicokinetics: general principles . . . . . . . . . . . . . . . . . . 388
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 427
7.4.2. Uptake . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 389
Selenium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 427
7.4.2.1. Bioaccumulation and biomagnification:
7.6.5.8. Marine mammals . . . . . . . . . . . . . . . . . . . . . . . . 427
general principles . . . . . . . . . . . . . . . . . . . . . . . . . 389
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 427
7.4.2.2. Terrestrial ecosystem: bioaccumulation/
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 428
biomagnification . . . . . . . . . . . . . . . . . . . . . . . . . 390
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 431
7.4.2.3. Freshwater ecosystem: bioaccumulation/
Selenium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 436
biomagnification . . . . . . . . . . . . . . . . . . . . . . . . . 390
7.7. Biological effects (acute, short-, and long-term toxicity;
7.4.2.4. Marine ecosystem: bioaccumulation/biomagnification 390
reproductive, physiological, and behavioral effects; etc.) . 437
7.4.3. Transport, biotransformation, and distribution . . . . . . . . 391
7.7.1. Effects on terrestrial ecosystems . . . . . . . . . . . . . . . . . . . . 437
7.4.4. Excretion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 393
7.7.2. Effects on freshwater ecosystems . . . . . . . . . . . . . . . . . . . . 437
7.4.5. Uptake, accumulation, and loss in biota . . . . . . . . . . . . . . 393
7.7.3. Effects on marine ecosystems . . . . . . . . . . . . . . . . . . . . . . 438
7.4.5.1. Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 394
Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 438
7.4.5.2. Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 394
Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 438
7.4.5.3. Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 395
Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 440
7.4.5.4. Selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 396
Selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . 440
7.5. Toxicological effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 396
7.8. Conclusions and recommendations . . . . . . . . . . . . . . . . . . . 440
7.5.1. Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 397
7.8.1. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 440
7.5.2. Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 397
7.8.1.1. Sources and transport of metals . . . . . . . . . . . . . . 440
7.5.3. Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 398
7.8.1.2. Arctic metal concentrations relative to
7.5.4. Selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 399
global background . . . . . . . . . . . . . . . . . . . . . . . . 442
7.6. Regional and circumpolar levels and trends of
7.8.1.3. Spatial trends within the Arctic . . . . . . . . . . . . . . 443
metal contamination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 400
7.8.1.4. Temporal trends within the Arctic . . . . . . . . . . . . 443
7.6.1. Atmosphere. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 400
7.8.1.5. Observed biological effect and health aspects
7.6.1.1. Air concentrations in the High Arctic . . . . . . . . . 400
attributable to metals . . . . . . . . . . . . . . . . . . . . . . 443
7.6.1.2. Concentrations of heavy metals in subarctic air . . 402
7.8.1.5.1. Observed biological effects . . . . . . . . . 443
7.6.1.3. Atmospheric deposition in the Arctic . . . . . . . . . . 402
7.8.1.5.2. Tissue burdens of metals relative to
7.6.2. Terrestrial ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . 405
national standards . . . . . . . . . . . . . . . . 443
7.6.2.1. Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 405
7.8.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 443
7.6.2.2. Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . 406
Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 443
7.6.2.3. Vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 406
7.6.2.4. Terrestrial birds . . . . . . . . . . . . . . . . . . . . . . . . . . 407
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 444
7.6.2.5. Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 409
Annex . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 453
374
AMAP Assessment Report
The ideal situation when comparing data for such a wide
7.0. Introduction
area and from diverse sources would be to have access to all
Heavy metals occur naturally in all ecosystems, but with
raw data (individual sample concentration values) and any
large variations in concentration. They are also released to
necessary supporting information (sample characterization,
the environment from many different anthropogenic sources.
biological information, full methodological descriptions).
This chapter identifies the metals that are relevant in a pol-
This was not possible for much of the data used in preparing
lution context and considers the chemical forms in which
this chapter, e.g., much of the data available in the published
they are found. It describes natural and man-made sources,
literature. For all the new AMAP data, however, appropriate
pathways, and levels of heavy metals in the atmosphere and
reporting procedures have been implemented. Notwith-
in the marine, freshwater, and terrestrial environments. Tox-
standing these limitations, a substantial understanding of the
icological characteristics of heavy metals, e.g., uptake, accu-
status of metals in the Arctic can be reached using the results
mulation, and effects in organisms, are also considered.
of already-published investigations.
A focus of this chapter is the description of concentra-
Methods used in the preparation of this chapter for the
tions of heavy metals found in the Arctic in terrestrial, fresh-
selection, handling, and assessment of data are described in
water, and marine ecosystems, because these data provide
the appropriate subsections.
the basis for assessing geographical differences and temporal
trends in levels of heavy metals. Although a substantial
amount of heavy metal data has been compiled, detailed
7.1. Physical /chemical characteristics
conclusions are difficult to make because 1) the area covered
7.1.1. Identification of metals to be considered
by the AMAP assessment is very large; 2) few areas have
been monitored on a regular basis (the best covered areas
Heavy metals can create adverse effects on environmental
have data for less than five sampling years, and temporal
and human health due to their toxicity and their bioaccumu-
trend sampling seldom covers up to two decades); 3) geo-
lation in various environmental compartments. A number of
graphical coverage of available sampling data is poor; and
studies have been carried out to assess the behavior of these
4) analytical results are often not completely comparable
pollutants in the environment (e.g., review in Pacyna et al.
due to sampling, analytical, and reporting differences.
1993a). The results of these studies are summarized in Table
These limitations can be severe. A number of publications
7·1 (after Nriagu 1984 and Pacyna and Winchester 1990).
have addressed the problem of what constitutes adequate
Environmental concentrations of many of the metals listed
biological sampling in connection with ecotoxicological in-
in this table are often higher than the concentrations expected
vestigations (Bignert et al. 1993, 1994, Olsson 1995). These
from their natural occurrence in terrestrial and aquatic envi-
works stress that substantial individual variation occurs in
ronments. Emissions of heavy metals, mostly on fine parti-
biological systems and that many samples have to be ana-
cles, during various human activities are the major cause of
lyzed before values are available which reliably describe bio-
these increased concentrations, resulting in alterations of
accumulation, seasonal variation, or spatial and temporal
geochemical cycles of these metals.
trends. According to Olsson (1995), from 11 to > 20 years of
The increase in concentration of a given metal, measured
annual sampling is needed to discover a 5% annual change
in a certain reference material, such as crustal rocks or soils,
in the concentrations of Cd, Pb, and Hg in the muscle and
in relation to a certain reference metal, such as Al, Ti, or Sc,
liver of Swedish reindeer.
can be defined as the enrichment factor of this metal (EF).
There are also problems linked to analytical quality. Over
Most often, metals are enriched on a local scale, but some
the last 25 years a number of different techniques (including
are enriched on regional and global scales. Regional scale is
differences in sample preservation and preparation, analysis,
often defined as continental (1000-2000 km), whereas glo-
and equipment) have been used to measure heavy metal con-
bal scale is usually regarded as intercontinental, e.g., North-
centrations. The available data are thus not always of uni-
ern Hemisphere. Episodes of long-range transport of pollu-
form quality. In a review of heavy metals in the Greenland
tants within air masses result in the enrichment of metal
marine environment (Dietz et al. 1996), it was evident that
concentrations far from source regions; the Arctic is a recep-
some of the older data were incorrect (too high) because
tor of such transport (Pacyna and Winchester 1990).
techniques had not been sufficiently tested at that time, par-
The environmental and health effects of heavy metals de-
ticularly for Pb and Cd. It should be stressed that this chap-
pend greatly upon on the mobility of each metal through en-
ter reports the data as they were published in the literature,
vironmental compartments and the critical pathways through
without adjusting for the different analytical techniques, ex-
which the metals reach the human body. Almost all metals
cept in cases where raw data or analytical information was
in Table 7·1 are either water/lipid soluble or volatile. Some
available to the authors (some data which were clearly erro-
metals are water/lipid soluble and volatile. Ingestion with
neous were omitted). It is obvious that considerable atten-
food is the major pathway for many metals entering humans;
tion must be directed at intercalibration of laboratories in
however, quantitative information is incomplete for many
the future if coherent data sets for circumpolar heavy metal
metals.
concentrations are to be obtained.
Finally, the degree of concern about human and envi-
Another factor that makes it difficult to compare existing
ronmental health varies with each metal. Some metals are
heavy metal data is the way in which data are reported, i.e.,
clearly toxic. Others are known to be essential micronu-
the statistical parameters reported, and the level of detail
trients for humans and animals. The true importance of
available in the reported data and its supporting informa-
some metals to human and animal health is not known due
tion. For example, much of the available data consists of
to incomplete information. The general indication of po-
arithmetic means, geometric means, or median values; statis-
tential health concern for some heavy metals is also noted
tical parameters which can differ considerably when calcu-
in Table 7·1.
lated for any given set of data. This complicates comparison
Taking into account the available information on the be-
of data from different sources. For biota, the species and tis-
havior and effects of heavy metals, most studies focus on
sues analyzed, age/size, sex, and year of sampling also differ
Hg, Cd, and Pb. Fewer studies target As, Cu, Cr, Ni, V, Se,
among the available data and make comparisons complex.
and Zn. These two groups of heavy metals have been pro-
Chapter 7 · Heavy Metals
375
Table 7·1. Perturbations of the geochemical cycles of trace metals by soci-
7.1.2. Speciation of metals in the environment
ety (Nriagu 1984, Pacyna and Winchester 1990).
Atmosphere
Scale of
perturbation a
Information on the chemical speciation of heavy metals
Health Critical
Ele-
Glo- Regio- Lo-
Most diagnostic
Mobil- con-
path-
emitted to the atmosphere is limited. For Hg, most of the
ment bal
nal
cal
environments b
ity c
cernd
way e
emissions from combustion of fuels occurs in the gaseous
phase. In the combustion zone, Hg present in coal or oth-
Pb
+
+
+
A, Sd, I, W, H, So
v, a
+
F, A f
V
+
+, c
+
A
g
(+)
A?
er fossil fuels evaporates in elemental form. Some of it is
As
+
+
+
A, Sd, So, W
v, s, a
+
A, W
then oxidized while in the flue gases. The oxidized forms
Sn
+
+
+
A, Sd, W
v, a
+8
F
of Hg can be retained in modern flue gas cleaning systems.
Zn
+
+
+
A, Sd, W, So
v, s
E
F
Mercury retained in fly ash (as well as in bottom ash) is
Cd
+
+
+
A, Sd, So, W
v, s
+
F
often disposed of on land, after which some is transported
Hg
+
+
+
A, Sd, Fish, So
v, a
+ h
F, (A)
Sb
+
+
+
A, Sd
v, s
(+)
F, W, A?
to the aquatic environment. While it is difficult to quantify
Cu
+
+
+
A, Sd, W, So
v, s
E
F?
these transport processes, Nriagu and Pacyna (1988) esti-
Ag
+
+
+
A, Sd, W
(v)
(+)
?
mated that Hg in coal fly ash and bottom ash contribute
Se
+
(+)
+
A
v, s, a
E
F
Ge
?
+
+
A, So, W?
v, s, a
(+) h
?
up to 40% of the direct releases of the element to the ter-
restrial environment. The emission generation process for
Ni
(+)
+
+
A, Sd
v9
E
F, W, A?
Hg during the incineration of wastes is similar to that
Cr
+
+
A, Sd, W, Gw
s, v i
E
W, F
B
(+)
+
A, Sd, Gw
v, s
E
W
during combustion of fossil fuels. However, more Hg in
K
(+)
+
A
s
E
F
the oxidized form is expected from incinerators due to
the higher content of chlorine in waste matter than in fos-
Pt
?
?
+
A, Sd
s
(+)
?
Pd
?
?
+
Sd
s
(+)
?
sil fuels.
Mo
?
?
+
A, W, So, Sd
s
E
F, W
The major chemical forms of As, Cd, and Pb created
Tl
?
?
+
Em, So
v, s
(+)
A, F?
by the main emission source categories are presented in
In
?
?
+
A, So, Em
v
(+)
?
Bi
?
?
+
A, So, Em
v
(+)
?
Table 7·2. The inorganic forms of As (particularly triva-
Be
?
?
+
A, So, Em
(+)
A
lent As), from sources including smelters or coal-fired
Ga
?
?
+
Em
v
(+)
?
power plants, dominate in the air over emission areas.
Te
?
?
(+)
So
v, a?
(+)
?
It is important to observe that inorganic compounds of
a. + : significant perturbation; (+) : possible perturbation; : no perturba-
arsenic generally are more toxic than organic arsenic com-
tion; ?: not enough information; c: enhanced due to mobilization of
pounds, and that the trivalent forms are more toxic than
crustal materials (soil, dust).
b. A: air; Sd: sediments (coastal, lake); So: soils; I: ice cores; W: surface
the other forms. Methylated forms of As are probably of
waters; Gw: groundwaters; H: humans; Em: emission studies (only
minor significance.
listed when little geochemical information is available).
Elemental Cd (Cd(0)) and its oxide are the predominant
c. v: volatile; s: soluble; r: soluble only under reducing conditions; a: mo-
bile as alkylated organometallic species; : not mobile.
chemical forms of the metal emitted from major sources.
d. +: toxic in excess; (+) : toxic, but little data available; E: essential, but
These two forms seem to be the most toxic Cd species, to-
toxic in excess.
gether with Cd chloride, which is found in emissions from
e. F : food; W: water; A: air;
f. Exposure through hand-to-mouth activity is critical for lead in children.
waste incineration. For Pb, inorganic forms are the most
g. Enriched relative to crustal abundance from fuel oil combustion (vana-
widely released chemical species, particularly Pb oxide, Pb
dium porphyrins).
chloride, and Pb sulfates.
h. Organometallic forms only.
i. Hexavalent form volatile and toxic, trivalent form essential.
The relative volatility of the chemical species presented in
Table 7·2 is as follows: elemental As, its trioxide and chlo-
posed by the United Nations Economic Commission for Eu-
ride, elemental Cd and its chloride, and elemental Pb are vol-
rope (UN ECE), as well as by international programs in-
atile; Pb chloride is intermediate; and the oxides of Cd and
volved in studies of the transport of pollutants to the North
Pb are non-volatile. The volatile species occur in a vapor
Sea and the Baltic Sea, as a priority list for emission reduc-
phase during emission generation, whereas the non-volatile
tion policies for heavy metals.
compounds are emitted largely as fly ash, even at higher
The Arctic region is a major receptor of heavy metals
temperatures.
generated in other regions of the Northern Hemisphere
(Rahn and Lowenthal 1984, Maenhaut et al. 1989, Barrie
Aquatic systems
1991, Shaw 1991b, Cheng et al. 1993). The aims of this
A complete understanding of chemical speciation is essential
chapter are 1) to assess emission sources and fluxes of heavy
for gaining a comprehensive understanding of the chemical
metals to and in the Arctic, 2) to describe the concentration
status of aquatic ecosystems. This is in turn essential for
of metals in various environmental compartments, 3) to as-
evaluating the risk to the health of the ecosystems and indi-
sess the degree of bioaccumulation and biomagnification of
viduals within them as a result of exposure to metals, and
heavy metals in Arctic biota, and 4) to assess the environ-
for being able to predict how changes in environmental pa-
mental effects of these pollutants in the Arctic. A major fo-
rameters will influence bioavailability, bioaccumulation, and
cus is placed on the priority heavy metals noted above, par-
the toxic effects of metals. An overview of metal speciation
ticularly Cd, Pb, Hg, and Se.
in aquatic systems is provided below. Greater detail is pro-
Table 7·2. Major chemical species created evolved during fossil fuel combustion and industrial processes. Me(O): elemental form of a given heavy metal.
Process
As
Cd
Pb
Coal combustion
As(0), As2O3, As2S3
Cd(0) , CdO, CdS
PbCl2, PbO, PbS, Pb
Oil combustion
As(0), As2O3, Organic arsines
Cd(0), CdO
PbO
Non-ferrous metal production
As2O3
CdO, CdS
PbO, PbSO4, PbO, PbSO4
Iron and steel manufacturing
CdO
PbO
Refuse incineration
As(0), As2O3, AsCl3
Cd(0), CdO, CdCl2
Pb(0), PbO, PbCl2
376
AMAP Assessment Report
vided in a number of good reviews (Sadiq 1992, Webb 1979,
Lead
Nriagu 1980, Furness and Rainbow 1990, Singer 1973,
Lead usually exists in natural waters as Pb(II). It reacts read-
NRCC 1979a, 1979b, 1979c, 1979d, 1981a, 1981b, Elinder
ily with some major anions (CO 2
2
3 , SO4 ) at pH values typ-
1984, Moore and Ramamoorthy 1987, Linnik and Nabiva-
ical of natural waters. Lead has a high complexing ability
nets 1986)
and forms stable complexes with S-, P-, O-, and N-contain-
The speciation of heavy metals in aquatic systems is
ing organic ligands which leads to its accumulation in live
controlled by a number of factors including ionic strength,
and dead aquatic biota. At low concentrations of soluble
pH, redox potential, presence of organic ligands, and tem-
organic ligands, Pb exists primarily in particulate form at
perature. Metals occur broadly in particulate and soluble
pH > 6. The proportion of particle-bound Pb to total Pb in
forms. Possible species include free aquated ions; complex
the world's river waters varies from 15 to 83% depending
ionic entities; inorganic ion-pairs and complexes; organic
on water composition. Like Hg, Pb can be microbiologically
complexes, chelates and compounds; metals bound to
methylated in bottom sediments.
high molecular weight organic materials; highly dispersed
The speciation of metals in aquatic systems has impor-
colloids; metals sorbed on colloids; precipitates; and met-
tant consequences for both bioaccumulation and toxicity.
als present in live and dead biota. Various species are in
In freshwater, free metal ions of Pb, Cd, and Hg are much
dynamic equilibrium with each other and thus the rel-
more abundant than in natural waters containing chloride
ative proportion of each species in an aqueous medium is
ion and other complexing anions; in general these divalent
determined by thermodynamic and kinetic relationships.
cations are considered to be more toxic than the soluble
Using Cd as an example, the following is a comparison
complexed forms of Pb, Cd, or Hg. Consequently, the toxic
of the dissolved chemical species present in `ideal' seawa-
effects of Pb, Cd, or Hg are frequently more pronounced in
ter and river water at 25°C. In seawater, over 90% of the
freshwater systems than in estuarine or marine systems, even
Cd is in the form of chloro complexes; of these CdCl2
through the concentrations of dissolved metals are higher in
represents more than 50% (CdCO
3 < Cd2+ < CdCl3 < CdCl+
the latter (complexation increases solubility, but usually re-
< CdCl2).
duces toxicity).
Equilibrium conditions in aquatic systems are often ex-
tremely complicated and difficult to measure; consequently,
chemical species in the system are often estimated from ther-
7.2. Sources of pollution
modynamic solubility relationships.
Heavy metals are emitted to the atmosphere from both nat-
Mercury
ural and anthropogenic sources. Very few of the sources that
Mercury exists in natural waters in three oxidation states:
directly result in the contamination of the Arctic environ-
Hg(0), Hg(I), and Hg(II). Hg(II) forms hydroxocomplexes
ment are located in the Arctic. Metals released to the envi-
even at low pH values. These complexes predominate among
ronment outside the Arctic are transported to the Arctic via
inorganic forms of Hg under natural conditions (pH > 6),
air currents, rivers, and ocean currents. This section reviews
even in the presence of considerable concentrations of chlo-
the sources and fluxes of heavy metals within the Arctic as
ride ions. Mercury can also form stable complexes with
well as their sources outside and pathways to the Arctic.
many organic ligands, especially those containing sulfur
(amino acids, oxycarbonic acids etc.) and high molecular
7.2.1. Natural sources
natural compounds like fulvic and humic acids. In natural
waters, Hg compounds are strongly bound to particulate
An accurate inventory of heavy metal sources and emissions
matter. In particulates, Hg is readily transported by river
to the atmosphere from natural processes is needed to make
and accumulates in bottom sediments.
a complete assessment of the extent of regional and global
In addition to hydrophobic organic complexes, Hg(II) can
pollution by heavy metals in the Arctic. It is generally pre-
form water- and lipid-soluble alkylderivative compounds,
sumed that the principal natural sources of heavy metals in-
primarily methylmercury. There are two pathways of methy-
clude wind-borne soil particles, volcanoes, seasalt spray, and
lation: microbiological and chemical. The later mechanism
wild forest fires (Nriagu 1989). Recent studies have shown,
usually involves natural organic substances, mostly fulvic
however, that particulate organic matter is the dominant
acids. Varshal and Buachidze (1983) report that 12 hours
component of atmospheric aerosols in non-urban areas (Tal-
after discharging Hg(II) into water containing high concen-
bot et al. 1988, Artaxo et al. 1988) and that over 60% of
trations of fulvic acid, up to 34% of the Hg is transformed
the airborne heavy metals in forested regions can be attri-
into alkyl-mercury compounds.
buted to aerosols of biogenic origin (Zoller 1983).
A global assessment of natural sources of atmospheric
Cadmium
heavy metals has been made by Nriagu (1989). A summary
Cadmium is usually present in the environment as Cd(II)
of these estimates is presented in Figure 7·1. Biogenic sources
and starts to be hydrolyzed at pH 9. The complexing ability
can account, on average, for over 50% of the Se, Hg, and
of Cd, even with humic and fulvic acids, is not as strong
Mo, and from 30 to 50% of the As, Cd, Cu, Mn, Pb, and Zn,
as other heavy metals; in natural waters it is predominant-
released annually to the atmosphere from natural sources.
ly present in dissolved (Cd(II)) form. Cadmium forms low
Volcanic emissions can account for 40-50% of the Cd and
solubility compounds with several anions (primarily car-
Hg and 20-40% of the As, Cr, Cu, Ni, Pb, and Sb released
bonate and phosphate) which control the concentration
annually from natural sources. Seasalt aerosols seem to ac-
of Cd(II) under natural conditions (pH 8-9). Low concen-
count for < 10% of atmospheric heavy metals from natural
trations of these anions favor the adsorption of Cd(II) by
sources. Finally, soil-derived dusts can account for over 50%
particulate matter, clay particles in particular, and Cd is
of the total Cr, Mn, and V emissions, as well as for 20-30%
transported in this form. Adsorption of Cd(II) by sedi-
of the Cu, Mo, Ni, Pb, Sb, and Zn released annually to the
ments increases with increasing pH; at pH > 7, nearly all
atmosphere. As the accuracy of emission estimates for nat-
Cd is in the adsorbed phase (Moore and Ramamoorthy
ural sources is low, these percentage contributions should be
1987).
considered as approximations only.
Chapter 7 · Heavy Metals
377
%
%
80
80
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
1.1 - 23.5
0.1 - 3.9
0.6 - 11.4
4.5 - 82.8
2.2 - 53.8
0 - 4.9
51.5 - 582
0.1 - 5.8
2.9 - 56.8
0.9 - 23.5
0.1 - 5.8
0.7 - 18.1
1.6 - 54.2
4.0 - 85.9
As
Cd
Co
Cr
Cu
Hg
Mn
Mo
Ni
Pb
Sb
Se
V
Zn
Windborne soil particles
Seasalt spray
Volcanoes
Wild forest fires
Biogenic processes
Numbers under columns : range of estimates (103 t/y).
Figure 7·1. Global emissions of trace metals to the atmosphere from natural sources (after Nriagu 1989). Numbers under the columns are the range of
estimates of the emissions in thousands of tonnes per year. The percentages shown by the bars are calculated using the maximum value of the range of
the total and individual source category estimates.
%
%
100
100
90
90
80
80
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
12 - 25.6
3.1 - 12.0
7.34 - 53.6
19.9 - 50.9
0.91 - 6.19
10.6 - 66
0.79 - 5.74
24.2 - 87.2
289 - 376
1.48 - 10.8
1.81 - 5.78
1.47 - 10.8
3.32 - 6.95
30.2 - 142
70.2 - 193
As
Cd
Cr
Cu
Hg
Mn
Mo
Ni
Pb
Sb
Se
Sn
Ti
V
Zn
Coal, oil and wood
Gasoline
Non-ferrous metal industry
Other industries and use
Waste incineration
Numbers under columns : range of estimates (103 t/y).
Figure 7·2. Global emissions of trace metals to the atmosphere at the beginning of the 1980s from anthropogenic sources (after Nriagu and Pacyna
1988). Numbers under the columns are the range of estimates of the emissions in thousands of tonnes per year. The percentages shown by the bars are
calculated using the maximum value of the range of the total and individual source category estimates.
The natural sources of heavy metals which influence the
transport of eroded dust from the deserts in Asia and Africa
freshwater, terrestrial, and marine environment are even
during dust storms. However, no quantitative assessment
more difficult to assess than the atmospheric sources. In gen-
has been made of how much of the eroded dust and attached
eral, soils and sediments tend to reflect the composition of
heavy metals is transported from the Asian and African des-
their parent material. Soils and sediments in mineralized
erts to the Arctic.
areas, therefore, usually have the highest concentrations of
the corresponding metals. For example, rocks with high Hg
7.2.2. Anthropogenic sources
content usually occur in areas of crustal instability where
volcanic and geothermal activity are high.
High temperature processes generate various heavy metals.
It is also very difficult to assess the extent to which emis-
These processes include coal and oil combustion in electric
sions from natural processes affect the contamination of the
power stations and heating and industrial plants, gasoline
Arctic environment. In general, fluxes from these processes
combustion, roasting and smelting of ores in non-ferrous
within the Arctic are regarded as less significant than an-
metal smelters, melting operations in ferrous foundries, re-
thropogenic releases, both within and outside the Arctic.
fuse incineration, and kiln operations in cement plants. The
However, very long range transport within air masses of soil
metals enter the atmosphere and the aquatic and terrestrial
particles from deserts in Asia and Africa to the High Arctic
ecosystems; virtually every industry discharges heavy metals
has been postulated by Pacyna and Ottar (1988). A series of
into these ecosystems. The assessment presented here is fo-
haze bands over Barrow, Alaska in April and May 1976
cused on the principal industrial and commercial users of
were found to consists of dust (Rahn et al. 1981). The bulk
raw materials and water, and on producers of solid wastes.
elemental composition of the particles was crustal or near-
crustal and their mass-median radius of about 2 m indi-
7.2.2.1. Sources and fluxes of atmospheric input
cated that they could have originated more than 5000 km
on a global scale
from Alaska. Trajectory analysis showed that these particles
could have passed over the arid and semi-arid regions of
The first quantitative worldwide estimate of the annual in-
eastern Asia during intense dust storms which had occurred
dustrial input of 16 heavy elements into air, soil, and water
there. This hypothesis has been confirmed by measurements
was published by Nriagu and Pacyna (1988). The summary
in the Norwegian Arctic (Pacyna and Ottar 1989) and in the
of the estimate of atmospheric emissions is presented in Fig-
Canadian Arctic (Welch et al. 1991). The origin and evolu-
ure 7·2. Pyrometallurgical processes in the primary non-fer-
tion of dust clouds in central Asia has recently received con-
rous metal industries are the major source of atmospheric
sideration. The existence of natural constituents in the Arc-
As, Cd, Cu, In, Sb, and Zn, and an important source of Pb
tic aerosol in central Asia was explained by long-range
and Se. Combustion of coal in electric power plants and in-
378
AMAP Assessment Report
%
%
100
100
90
90
80
80
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
31
8.9
74
63
6.1
355
6.3
86
344
5.9
16
114
177
As
Cd
Cr
Cu
Hg
Mn
Mo
Ni
Pb
Sb
Se
V
Zn
Anthropogenic sources
Natural sources
Numbers under columns : median values of the estimates (103 t/y).
Figure 7·3. Comparison of global emissions of trace metals to the atmosphere from natural and anthropogenic sources in 1983. Numbers under the
columns are the median values of estimates of total emissions in thousands of tonnes per year. The percentages shown by the bars are calculated from
the median values of the ranges of the estimates for natural and anthropogenic sources.
%
%
80
80
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
17 536
74 325
5 378
69 553
26 789
15 066
208 647
Africa
Asia
Australia
Europe
North America
South America
All regions
Gasoline combustion
Non-ferrous metal industry
Cement production
Fossil fuel combustion
Waste incineration
Iron and steel
Numbers under the columns : total emissions (t).
Figure 7·4. Global emissions of Pb to the atmosphere from various sources and source regions in 1989. Numbers under the columns are maximum esti-
mates of the total emissions in tonnes.
dustrial, commercial, and residential burners is the major
Combustion of leaded gasoline is still the major source of
source of anthropogenic Hg, Mo, and Se and a significant
Pb. Chromium and Mn are derived primarily from the iron
source of As, Cr, Mn, Sb, and Ti. Combustion of oil for the
and steel industry. Little information is available on the emis-
same purpose is the most important source of V and Ni.
sion of heavy metals from various diffuse (fugitive) sources.
A comparison of the median values of worldwide emis-
sions of heavy metals from natural and anthropogenic sources
(Figure 7·3) suggests that human activities generate emissions
of heavy metals that exceed those from natural sources. There-
fore, anthropogenic emissions result in significant alterations
of the natural biogeochemical cycling of many heavy metals
in the global environment.
Recently, a revision of the global emission inventory of Pb
was prepared for the reference year 1989 (Pacyna et al. 1993b)
as a part of the Global Emission Inventory Activities (GEIA)
operated within the IGBP International Global Atmospheric
Chemistry (IGAC) program. The results show that in 1989
the maximum emission was about 209 000 tonnes of Pb, of
which 62% came from gasoline combustion, followed by
26% from non-ferrous metal production (Pacyna et al. 1995).
The summary of the results showing the maximum emission
estimates is presented in Figure 7·4. One-third of the total
estimated emissions of the element originates in Asia and
Pb, t/y
Europe. The spatial distribution of these emissions within
<1
the 150
150 km EMEP grid system for the area north of
1 - 10
latitude 50°N is presented in Figure 7·5.
10 - 100
100 - 1 000
An estimate of global anthropogenic emissions of Hg has
>1 000
also been completed for AMAP (Pacyna and Pacyna 1996).
The spatial distribution of these emissions within a 1°
1°
No reported emissions
grid is presented in Figure 7·6, and the contributions from
Figure 7·5. Spatial distribution of emissions of Pb within the 150 km
different continents and different sources are presented in
150 km EMEP grid system for the area north of latitude 50°N in 1989.
Figure 7·7. These preliminary data suggest that between
(Source of data: Norwegian Meteorological Institute, after Pacyna et al.
1993b).
1300 and 2150 tonnes of Hg are emitted annually to the at-
Chapter 7 · Heavy Metals
379
Hg, t/y
No reported emissions
< 0.1
0.1 - 0.5
0.5 - 2
2 - 5
5 - 22
Figure 7·6. Spatial distribution of global emissions of Hg in 1990 within a 1°
1° grid. The total emission inventory is 2144 tonnes Hg. (Source of data:
Jozef Pacyna pers. comm., Canadian Global Emmissions Interpretation Centre (CGEIC)).
%
Africa
%
Asia
%
Australia
%
Europe
%
North America
%
South America
100
100
100
100
100
100
80
80
80
80
80
80
60
60
60
60
60
60
40
40
40
40
40
40
20
20
20
20
20
20
0
0
0
0
0
0
Coal combustion
Caustic soda production
Gold production
Waste disposal
Other sources
Figure 7·7. Contributions from different continents and from different sources to the global emissions of Hg to the atmosphere.
mosphere at the present time. The major sources of these
Ni), coal-burning power plants (As, Hg, and Se in particu-
emissions are combustion of coal to produce electricity and
lar), non-ferrous metal smelters (Cd, Ni, Pb, and Se), iron
heat (60%), followed by gold production and waste disposal.
and steel plants (Cr, Mo, Sb, and Zn), and the dumping of
The above-mentioned studies on global releases of heavy
sewage sludge (As, Mn, and Pb). The atmosphere is the ma-
metals are based on a number of research projects conducted
jor route of Pb entry into natural waters and also accounts
to estimate the atmospheric emissions of heavy metals in Eu-
for over 40% of the V loading. The results of the worldwide
rope, North America, and Asia. Emissions from these areas
assessment of anthropogenic inputs of 13 heavy metals into
relevant to the Arctic are reviewed below.
the aquatic ecosystem is presented in Figure 7·8 (Nriagu and
Pacyna 1988).
A comparison of data in Figures 7·2 and 7·8 indicates
7.2.2.2. Sources and fluxes of aquatic input
that for most of the heavy metals, the annual anthropogenic
on a global scale
inputs into water exceed the quantities emitted to the atmos-
The major sources of anthropogenic heavy metal contamina-
phere. If it is assumed that only 25% of the industrial efflu-
tion of aquatic ecosystems (including the ocean) include do-
ents are discharged into lakes and rivers, the average con-
mestic wastewater effluents (especially As, Cr, Cu, Mn, and
centrations in these waters should reach levels several-fold
%
%
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
11.6 - 70.3
2.1 - 16.3
45.6 - 239
34.7 - 191
0.2 - 8.8
109 - 415
1.8 - 21.2
33.1 - 194
97.2 - 277
3.6 - 32.2
10.1 - 71.9
2.1 - 20.8
77.5 - 395
As
Cd
Cr
Cu
Hg
Mn
Mo
Ni
Pb
Sb
Se
V
Zn
Mining, smelting
Manufacturing
Waste disposal
Steam electric generation
Atmospheric deposition
and refining
processes
Numbers under columns : range of estimates (103 t/y).
Figure 7·8. Global anthropogenic inputs of trace metals to aquatic ecosystems at the beginning of the 1980s (after Nriagu and Pacyna 1988). Numbers
under the columns are the range of estimates of the inputs in thousands of tonnes per year. The percentages shown by the bars are calculated using the
maximum value of the range of the total and individual source category estimates.
380
AMAP Assessment Report
%
%
70
70
60
60
50
50
40
40
30
30
20
20
10
10
0
0
52.4 - 112
5.6 - 37.7
485 - 1310
542 -1400
1.6 - 15
706 - 2630
29.8 - 145
93.3 - 494
479 - 1040
4.8 - 47.5
6 - 76.5
21.4 - 138
689 - 1950
As
Cd
Cr
Cu
Hg
Mn
Mo
Ni
Pb
Sb
Se
V
Zn
Waste disposal
Wastage of commercial products
Coal and bottom fly ash
Fertilizer
Peat (agricultural and fuel use)
Atmospheric deposition
Numbers under columns : range of estimates (103 t/y).
Figure 7·9. Global anthropogenic inputs of trace metals to soils at the beginning of the 1980s (after Nriagu and Pacyna 1988). Numbers under the
columns are the range of estimates of the inputs in thousands of tonnes per year. The percentages shown by the bars are calculated using the maximum
value of the range of the total and individual source category estimates.
higher than those in unpolluted lakes and rivers. In other
These include re-emissions of previously deposited Hg as
words, the current rate of worldwide industrial inputs great-
well as emissions resulting from discharge into water bodies
ly exceeds the baseline burdens of heavy metals in the aver-
and from contaminated soils. Hence, it is more appropriate
age lake and river. Most of the effluent discharges occur in
to differentiate between pre-industrial and post-industrial
Europe (including Russia), North America, and some Asian
diffuse sources (Lindqvist 1991) than between natural and
countries, implying that the contamination of the freshwater
anthropogenic re-emission.
resources in these regions may be much more severe than is
Only sparse information is available on re-emission of
generally realized.
other heavy metals from soils and water surfaces. Oceans
can be an important source of Se emissions to the atmos-
phere on a global scale, contributing as much as 25% to
7.2.2.3. Terrestrial input and output
the total emissions of the element (Nriagu 1989).
of heavy metals on a global scale
The first quantitative assessment of worldwide fluxes of
7.2.3. Emission inventories for sources
heavy metals into soils was prepared by Nriagu and Pacyna
within and outside the Arctic
(1988) for the reference year 1983. A summary of this work
is presented in Figure 7·9. The estimates suggest that soils
Obviously, only a part of worldwide emissions is responsible
are receiving large quantities of heavy metals from disposal
for the contamination of the Arctic environment by heavy
of a variety of industrial wastes. The two principal sources
metals. It is critical to identify sources important to the Arc-
of heavy metals in soils worldwide are the disposal of ash
tic and to quantify the amount of emissions from these
residues from coal combustion and the general breakdown
sources that reaches the Arctic region. The results of source-
and weathering of commercial products on land. Urban re-
receptor studies, summarized in Pacyna (1991), indicate
fuse represents an important source of Cu, Hg, Pb, and Zn
that emissions from sources in Eurasia contribute more
with notable contributions of Cd, Pb, and V also coming via
than half of the air pollution measured in the Arctic. The
the atmosphere. The large volumes of wastes associated with
major source regions include the Urals, the Kola Peninsula,
animal husbandry, logging, and agricultural and food pro-
the Norilsk area, and the industrial regions in Central and
duction can significantly affect the heavy metal budget of
Eastern Europe (Rahn and McCaffrey 1980, Rahn and
many soils. Although municipal sewage sludge may not be a
Lowenthal 1984).
particularly important source on a global scale, it can be one
The contributions of European and North American emis-
of the most important sources of metal contamination of
sions to Arctic air pollution seem to be smaller than the con-
soils on a local scale.
tribution from the Russian sources. European and North
If the estimated metal inputs are distributed uniformly
American emissions are, however, major contributors to the
over the cultivated land area, the annual rates of metal ap-
plication are not very significant because of the large back-
%
%
ground reservoir of heavy metals. Nevertheless, each soil
80
80
70
70
has a limited retention capacity for heavy metals and there
60
60
is growing concern that, at the current rate of anthropo-
50
50
genic input, many soils in various parts of the world (e.g.,
40
40
central Europe and Japan) either have become or will soon
30
30
become overloaded with heavy metals (Kabata-Pendias
20
20
1984, Asami 1988).
10
10
0
0
Soils, like waters, can be a source of atmospheric conta-
2 582
896
726
58 130
32 947
mination of some heavy metals, particularly Hg. These emis-
As
Cd
Hg
Pb
Zn
sions result from various out-gassing of Hg laden rock and
Stationary fuel combustion
Gasoline combustion
from volatilization of Hg from soils, vegetation, and water
bodies. Current data suggest that these emissions of Hg are
Non-ferrous metal industry
Other sources
of the same order as emissions from anthropogenic sources
(Pacyna and Keeler 1994).
Iron and steel production
Numbers under columns : emission (t/y).
It should be noted that emissions from natural sources
Figure 7·10. Emissions of selected heavy metals to the atmosphere in
are difficult to distinguish from so-called secondary emis-
Europe at the beginning of the 1990s. Numbers under the columns are
sions and diffusive re-emissions from anthropogenic sources.
emissions in tonnes per year.
Chapter 7 · Heavy Metals
381
64
70
142
9.1
Avonmouth
25
5.6
9.3 9
55.7
5
Penarroya
Bergbau
Konstantinovka
30
37
7
5
Vladikavkaz
Asturiana
24
Bor-Rudarsko
16
28
2.5
6
14.5
Trepce
Plovdiv
Porto Vesme
Emissions (t/y) : As
Cd
Figure 7·11. The ten major point sources, and their emissions of As and Cd to the atmosphere, in Europe at the beginning of the 1990s. Numbers above
the columns are emissions in tonnes per year.
contamination of the subarctic regions, such as northern
consumption of fuels in Eastern Europe in recent years is
Scandinavia (European emissions) and the northern part of
an important factor which has caused the decrease of Hg
Canada (North American emissions).
emissions.
These sources, located outside the Arctic, are important
Pacyna et al. (1991) project that by using best available
in discussing the contamination of the High Arctic environ-
technology, As and Cd emissions in Europe should decrease
ment, which remains largely unaffected by local industrial
by a factor of 3 and 2, respectively, by the year 2000. A prompt
activities. It is necessary, therefore, to review the emissions
switch to unleaded gasoline should reduce Pb emissions in
from these outside sources in order to assess quantitatively
Europe by a factor of 4-10.
their contribution to Arctic contamination.
Pursuant to the requirements of the 1990 US Clean Air
Act Amendment, an interim toxic emission inventory has
7.2.3.1. Atmospheric emissions from sources
outside the Arctic and their trends
Cd
Tonnes
The first attempt to estimate atmospheric emissions of
3 000
heavy metals from anthropogenic sources in Europe
2 500
was completed at the beginning of the 1980s (Pacyna
2 000
1984). This European survey has since been updated,
1 500
completed, and emission gridded (Axenfeld et al. 1992,
1 000
a review by Pacyna 1994). National emission inventories
500
0
have only recently become available in the European
1955
1960
1965
1970
1975
1980
1985
1990
countries. These emission data, together with the inter-
national expert estimates, were used to compile current
Pb
European emission estimates for As, Cd, Hg, Pb, and Zn
Tonnes
(Figure 7·10). A spatial distribution of the heavy metal
160 000
emission estimates in Europe is available within the EMEP
120 000
150 km by 150 km grid system (Axenfeld et al. 1992).
80 000
The ten major European point sources of heavy metal
40 000
emissions to the air at the beginning of the 1990s are
shown in Figure 7·11.
0
1955
1960
1965
1970
1975
1980
1985
1990
Changes of heavy metal emissions to the atmosphere
from sources in Europe from the 1950s until present have
Zn
also been studied. The results are presented in Figure 7·12
Tonnes
(Olendrzynski et al. 1995). A decreasing trend of Hg levels
120 000
in atmospheric deposition in Scandinavia has been observed
100 000
during the last few years (Munthe et al. 1994). This trend
80 000
was related to a possible decline of Hg emissions, particu-
60 000
larly in Central and Eastern Europe. These emissions can be
40 000
expected to decrease by up to 30%. The economic decline in
20 000
this part of Europe at the beginning of the 1990s, related to
0
the transition from centrally planned economies to market
1955
1960
1965
1970
1975
1980
1985
1990
oriented ones, was suggested as the major reason for the
Figure 7·12. Changes with time in European atmospheric emissions of Cd,
possible changes in Hg emissions. In addition, the lower
Pb and Zn.
382
AMAP Assessment Report
%
%
80
80
70
70
As
Kola
60
60
Peninsula
165
50
50
St. Petersburg
area
40
40
4
Pechora Basin
30
30
11.5
Yakutsk
Donetsk
Moscow
20
20
area
area
area
Norilsk area
6.5
63
16
10
10
246
0
0
Urals
2 332
307
5 405
310
Baïkal / Irkutsk
551
area
As
Cd
Pb
Hg
Kuznetsk
55
area
Fossil fuel combustion
429
Tonnes
Metal application
Fergana area
Caucasus
980
1 000
253
Industrial processes
Solid waste disposal
500
200
50
Numbers under columns : emissions (t/y).
Figure 7·13. Estimates of emissions of As, Cd, Pb and Hg to the atmos-
5
phere from major source categories in the United States. Numbers under
the columns are emissions in tonnes per year. (Source of data: As and Pb,
Voldner and Smith 1989; Cd, US EPA 1993a; Hg, US EPA 1993b).
Cd
%
%
Kola
100
100
Peninsula
St. Petersburg
90
90
29
area
80
80
2
Pechora
Norilsk area
70
70
Basin
Moscow
Donetsk
3.5
26
60
60
area
area
Yakutsk
6
40
50
50
area
Urals
2
145
40
40
30
30
Baïkal / Irkutsk
area
20
20
Kuznetsk area
14
262
10
10
Tonnes
0
0
Caucasus
Fergana area
300
471
322
69
1 688
31
845
54
274
150
As
Cd
Cr
Cu
Hg
Pb
50
Fuel combustion (stationary
5
sources)
Solid waste incineration
Industrial processes
Miscellaneous sources
Transportation
Ni
Numbers under columns : emissions (t/y).
Kola
Figure 7·14. Estimates of emissions of As, Cd, Cr, Cu, Hg and Pb to the
Peninsula
St. Petersburg
645
atmosphere from major source categories in Canada in 1982 (after Jaques
area
1987). Numbers under the columns are emissions in tonnes per year.
80
Norilsk area
Pechora
935
Donetsk
Moscow
been developed for the continental United States. Emission
Basin
Yakutsk
area
area
73
area
506
300
estimates for As, Pb (both after Voldner and Smith 1989),
30
Urals
Cd (after US EPA 1993a), and Hg (after US EPA 1993b)
1620
from major source categories in the United States are pre-
Kuznetsk area
Baïkal / Irkutsk
138
area
sented in Figure 7·13.
50
Environment Canada has initiated several projects on
Caucasus
Tonnes
emission inventory development for heavy metals in Cana-
75
1 600
Fergana area
210
900
da. In Figure 7·14, the 1982 emissions of As, Cd, Cr, Cu,
500
Hg, and Pb from major source categories are presented on
200
30
the basis of data from Jacques (1987). Emission estimates
for Hg and Pb have been revised to account for major
changes in consumer patterns in recent years.
Zn
It is believed that the reliability of emission data from Eu-
rope and North America decreases in the following order:
Kola
St. Petersburg
Peninsula
area
180
Pb > Hg and Cd > remaining heavy metals.
20
Pechora
Moscow
An accuracy of < 25% was suggested for the emission esti-
Donetsk
Basin
Norilsk area
area
area
56
262
Yakutsk
74
mates of Pb, 50% or less for Cd and Hg, and 100% for the
area
2520
26
rest of the metals (Pacyna 1994).
Urals
3920
Baïkal / Irkutsk
area
7.2.3.2. Atmospheric emissions
Kuznetsk area
88
8830
from sources within the Arctic
Tonnes
Caucasus
Fergana area
8 500
266
4550
In addition to outside sources, there are also sources of
4 500
2 500
heavy metals within the Arctic. Combustion of fossil fuels
1 000
300
to produce electricity and heat is one of the major source
categories present in the region, followed by industrial
Figure 7·15. Emissions of As, Cd, Ni and Zn to the atmosphere from
processes in the Russian Arctic.
major sources in the former Soviet Union in 1979/80. (After NILU 1984).
Chapter 7 · Heavy Metals
383
7.2.3.3. Aquatic emissions
from sources outside the Arctic
Compared with atmospheric emissions, much less informa-
tion is available on emission inventories reporting discharges
Norilsk
of heavy metals to the aquatic environment. In one of very
few approaches in Europe, discharges of Cu, Pb, Ni, and Zn
from major point sources were estimated for the Commission
KrasnouralskSredneuralsk
Kushtym
of the European Communities (EC) (Daamen et al. 1990).
Verkhnaya Dyshma
Aztemowski
Balkhashski Gorno
Pyshma
These estimates are presented in Figure 7·17. Ship building
Mednogorsk
Pb, t/y
(including ship maintenance) is estimated to be the major in-
Irtych
200
dustry discharging Cu to the aquatic environment in Europe.
Dzhezkazganski
Gorno
400
Nickel discharges in the EC region originate mainly from
manufacturing of electrical equipment and the secondary
800
Tashkent
transformation of metals. The manufacture of ceramics is by
far the most important industry discharging Pb to the aquat-
Figure 7·16. Major point sources (non-ferrous metal smelters) of Pb to the
ic environment. Finally, Zn discharges in the EC region are
atmosphere in the Urals and the Asian part of the former Soviet Union.
mainly due to the manufacturing of basic industrial chemi-
(After Pacyna et al. 1993b).
cals. Daamen et al. (1990) conclude that discharges of heavy
Preliminary estimates of atmospheric emissions of As,
Cd, Cu, Cr, Mn, Ni, Pb, Sb, Se, V, and Zn for major source
Others
Shipbuilding
Copper
regions in the former Soviet Union, including the areas out-
23%
side of the Arctic, have been prepared by NILU (1984) on the
28%
basis of emission factors and statistical data for the reference
year 1979/80; data for selected metals are shown in Figure
8%
Textile finishing
7·15. Emissions from non-ferrous and ferrous metal produc-
16%
9%
Secondary
tion, fossil fuel combustion, and gasoline combustion were
Manufacture of
16%
ceramics
transformation
estimated to dominate the total emissions in those regions en-
of metals
tirely or partly located within the Arctic, such as the Kola Pen-
Manufacture of basic
insula, the Norilsk area, and the northern Urals. The inven-
industrial chemicals
tory of sources in the European part of the former Soviet Un-
ion is included in the European emission survey noted above.
Nickel
Others
A list of major point sources of heavy metal emissions to the
Manufacture of electrical
Non-ferrous
13%
equipement
atmosphere in the Urals and the Asian part of the former So-
metal industry
29%
viet Union is presented in Figure 7·16 (Pacyna et al. 1993b).
7%
Manufacture of
vegetable oil and fat
Sivertsen et al. (1992, 1994) studied the level of air
3%
4%
pollution in 1989 in the border areas of Norway and Rus-
Manufacture of
motor vehicles
sia related to the Cu and Ni emissions from the sources on
16%
Secondary
the Kola Peninsula. With respect to emissions of heavy
28%
Manufacture of basic
transformation
industrial chemicals
metals, there are two major source regions on the Kola
of metals
Peninsula: the Pechenganikel industrial complex consisting
of the Nikel and Zapolyarnyy Cu-Ni smelters, and the
Others
Lead
Severonikel smelter complex. The emissions of Cu and Ni
Iron and
steel industry
in the Pechenganikel smelter complex are estimated to be
9%
6%
Non-ferrous
approximately 310 and 510 tonnes, respectively. How-
metal industry
7%
Manufacture of ceramics
ever, very recent information (e.g., Pozniakov 1993, Ly-
41%
Manufacture of basic
angusova 1990) suggests that actual emissions could be
6%
industrial chemicals
about one order of magnitude higher. By contrast, the of-
7%
ficial Russian data place the 1994 emissions from Nikel
Fertilizer industry
12%
12%
and Zapolyarnyy at about 163 tonnes of Cu and 297 ton-
Secondary transformation
nes of Ni (CENR 1995).
of metals
Titanium dioxide
During the 1980s, the Severonikel smelter complex be-
industry
came the largest Ni producer in the world with an annual
output of 140 000 tonnes. Preliminary estimates of Cu and
Ni emissions to the atmosphere from this source are approxi-
Others
Manufacture of basic
Zinc
industrial chemicals
mately 3000 and 2700 tonnes, respectively (Pozniakov 1993).
24%
27%
These emissions appear high, particularly when compared
with the historical estimates of emissions from the Pechenga-
4%
Non-ferrous metal industry
nikel complex. The official Russian data place the 1994 emis-
7%
Iron and steel industry
sions from Severonikel at about 934 tonnes of Cu and 1619
18%
Secondary transformation
3%
of metals
tonnes of Ni (CENR 1995). The high variability among vari-
8%
Printing industry
9%
ous estimates for emission, however, emphasizes the need to
Manufacture of
ceramics
Titanium dioxide
verify the information on heavy metal emissions on the Kola
industry
Peninsula because these emissions contribute substantially to
Figure 7·17. The relative importance of different branches of industry in
the contamination of the Norwegian Arctic and perhaps to
discharges of Cu, Ni, Pb and Zn to the aquatic environment from major
that of a broader area.
point sources in the European Community (EC). (After Daamen et al. 1990).
384
AMAP Assessment Report
metals from the iron and steel industry, non-ferrous metal
In 1988, the UN Economic Commission for Europe (UN
manufacturing, and mining in the CEC region are under-
ECE) established a Task Force on Heavy Metals with the
estimated in the above-mentioned inventory due to lack of
aim of providing a state-of-the art report on heavy metals,
information.
examining the following issues: emission inventories, atmos-
pheric dispersion and deposition, analytical problems, and
abatement techniques including their economic aspects (UN
7.2.3.4. Aquatic emissions from sources within the Arctic
ECE 1994). The ultimate goal of the Task Force is to pre-
In Greenland, mining has been an important source of local
pare a substantiation document for a protocol on the reduc-
heavy metal pollution of the sea. Asmund et al. (1991) have
tion of heavy metal emissions in the ECE region as well as
made an estimate of quantities of Zn, Cd, and Pb released
in the United States and Canada.
to the sea from two mines in Greenland. One of these mines,
the Black Angel Pb-Zn mine in Uuummannaq, West Green-
land, operated from 1973 to 1990. During production, it
7.3. Special issues of pathways of metals
was estimated that 8-16 tonnes of Zn, 6-12 tonnes of Pb,
and 50-120 kg of Cd were released annually to seawater.
Pathways of metals reaching the Arctic are studied with the
Dumping of waste rock at the time the mine was closed in
use of transport models. This section presents models simu-
1990 added about 10 tonnes of Zn, 1 tonne of Pb, and 60
lating long-range transport of these compounds to the Arctic.
kg of Cd to this total (Asmund 1992a, 1992b). After mine
closure, the amounts released have declined drastically, to
7.3.1. Atmospheric transport
10% or less of the amounts measured during production
(Asmund 1992a, 1992b). At a closed cryolite mine in Ivit-
Following release into the atmosphere, heavy metals can be
tuut in south Greenland, Asmund et al. (1991) estimate an
either deposited in the vicinity of the emission source or sub-
annual input to the sea of 1-2.5 tonnes of Zn, 0.4-1 tonnes
ject to long-range transport via air masses. In most cases
of Pb, and 4-10 kg of Cd from deposited waste rock in the
(except Hg and to some extent Se), emission of heavy metals
intertidal zone. No estimates are available from a closed Pb
occurs on particles. The size of emitted particles containing
mine at Mestersvig in East Greenland, which operated from
heavy metals, as well as the temperature and speed of ex-
1956 to 1963, but environmental studies of seaweed, mus-
haust gases and the height of the emission source, are the
sels, and fish indicate that significant amounts of Pb and Zn
major factors influencing the relative proportion of metals
have entered the marine environment (Johansen et al. 1985).
transported locally over long distances.
A comparison of the Greenland mines with the Pb-Zn
During the past two decades, there have been several ap-
mine at Nanisivik in the Canadian Arctic shows that the
plications of various long-range transport models to study
Canadian mine has impacted the sea to a much lesser extent
the origin of atmospheric heavy metals measured in large in-
(Asmund et al. 1991). Information on the releases of metals
dustrial and residential areas, such as the entire European
from the Red Dog Pb-Zn mine in Alaska and their impact
continent, and in remote locations, such as the northern
on the environment is not yet available.
parts of Scandinavia and some parts of the Arctic. One focus
The major mining and metallurgical activities on the Kola
has been assessing the degree to which atmospheric deposi-
Peninsula and in the Norilsk region are sources of heavy
tion of these pollutants contributes to the contamination of
metals into the aquatic environment (e.g., Falk-Petersen et
certain regions, including the North Sea, the Baltic Sea, and
al. 1992). For example, the Severonikel smelter in Monche-
the Arctic. Despite intense interest in using transport models
gorsk discharged approximately 24 million m3 of waste-
in relation to PARCOM and HELCOM activities (studies of
water containing 54 tonnes of Ni during 1993 (NEFCO
the transport and deposition of pollutants to the North Sea
1995). The waters around Novaya Zemlya are important
and the Baltic Sea, respectively) and other projects concern-
dumping areas for various wastes, including toxic wastes.
ing deposition over the entire European continent and over
However, no quantitative assessment exists for the dumping
the drainage basins of the rivers Rhine and Elbe, only a few
of heavy metals contained in these wastes within this region.
operational models validated through comparison with ex-
Heavy metals are also transported by the large rivers, such
perimental data are currently available in Europe. These
as the North Dvina, Pechora, Ob, Yenisey, and Lena, con-
models have been designed to simulate long-range and long-
tributing to the contamination of the Arctic environment.
term transport of inert, particle-bound heavy metals and in-
clude 1) a combined trajectory-climatologic approach (Al-
camo et al. 1992), 2) statistical approaches of a Gaussian
7.2.3.5. International agreements on emission reduction
plume model and a trajectory model (van Jaarsveld et al.
A number of international efforts have been initiated to try
1986), 3) an EMEP-type Lagrangian trajectory model (Pe-
to reduce heavy metal emissions to the environment. These
tersen et al. 1989, Petersen and Kruger 1992), and 4) a Eu-
include:
lerian model (Galperin et al. 1994a).
1. Paris Commission (PARCOM) Decisions. A variety of
actions have been agreed to through the auspices of the
7.3.1.1. Dispersion models to study the impact
Paris Commission to reduce heavy metal emissions to the
of sources outside the Arctic
air and marine environment.
The first model of long-range transport of heavy metals in
2. Third North Sea Conference 1990. Many countries un-
Europe (Pacyna et al. 1984) was presented soon after the
dertook to implement reduction of discharges of heavy
first European emission survey was completed. The model
metals.
was similar to that used in the OECD study on long-range
3. Organization for Economic Co-operation and Develop-
transport of sulfur compounds (OECD 1979). The model
ment (OECD). The OECD is preparing a series of docu-
results were used to verify emission data through a compari-
ments examining risk reduction measures for a variety of
son of measured and calculated concentrations of heavy
substances under the OECD Chemical Programme, in-
metals in ambient air. Later, the model was modified and
cluding documents for heavy metals.
used to study the long-range transport of heavy metals to
Chapter 7 · Heavy Metals
385
to study long-range transport of Pb, Cd, Zn, and As from
Winter
sources in Eurasia and North America to the Arctic (Galpe-
rin et al. 1994b). It was concluded that the simulation of the
deposition of studied metals in the Arctic demonstrated an
appreciable impact of emissions from sources in these regions.
The impact of emissions from Eurasian sources on the
quality of air in Barrow, Alaska during 1985-1992 was also
studied using an isentropic air trajectory model, developed
by Harris and Kahl (1994). During the Arctic haze season
(wintertime), trajectories suggested that the transport of pol-
lution from north central Russia occurs near the surface,
whereas that from northern Europe occurs at higher altitudes.
In a more quantitative approach, a long-range aerial trans-
port Lagrangian model of Olson and Oikawa (1989) was
used to compute the annual flux of Sb, As, Cd, Pb, Zn, and
V into the Arctic from Eurasia (Akeredolu et al. 1994). Com-
parison of the model-predicted concentrations of heavy met-
als with a set of limited observations at existing sampling
stations close to the Arctic Circle showed agreement within
a factor of 2 to 3.
Pb, ng/m3
Recently, a model simulation of Pb transport to the Arctic
< 0.1
has been performed for AMAP on the basis of the global
0.1 - 0.25
> 4
emission inventory mentioned earlier (Figure 7·5, Tarrason
Pb, ng/m
0.25 - 0.5
3
0.5 - 1
1996). Air concentrations (Figure 7·18) and deposition (Fig-
Summer
< 0.1
1 - 2
ure 7·19) of this metal within the Arctic were estimated for
0.1 - 0.25
2 - 4
the reference year 1988. In general, the results of this simu-
0.25 - 0.5
> 4
lation confirm the impact of emission sources in Eurasia on
0.5 - 1
1 - 2
the contamination of Arctic air.
2 - 4
The models used in Europe, their attributes, and main
> 4
area of application have been reviewed by Petersen (1993).
Summer
A similar comparison for the North American models has
been made by Voldner (in Pacyna et al. 1993a).
In summary, various modeling studies on the origin of
Arctic air pollution support the conclusion that during win-
Figure 7·18. Averaged upper (3000 m) air concentrations of Pb in winter
(December-February) and summer (June-August) as modeled by the up-
dated (1996) hemispheric EMEP transport model in a simulation for the
reference year 1988. The gridded Pb emissions data used in the model
simulation are shown in Figure 7·5. (Source of data: Norwegian Meteoro-
logical Institute, Leonor Tarrasón pers. comm.).
remote sites in Scandinavia (Pacyna et al. 1989) and finally
to the Norwegian Arctic. It was concluded that the emission
estimates for As, Sb, V, and Zn from anthropogenic sources
in Europe (as applied in the model) can be related to the
concentrations measured in the air over Scandinavia and the
Pb, mg/m2
Arctic, and that the agreement between measured and mod-
eled concentrations was within a factor of two.
< 0.05
A Eulerian model has been developed at the Meteorolo-
0.05 - 0.1
0.1 - 0.5
gical Synthesizing Centre - East (MSC-E) of EMEP (Galperin
0.5 - 1
et al. 1994a). The model is used to calculate source-receptor
1 - 10
> 10
relationships for Pb, Cd, As, and Zn in the UN ECE region.
An important improvement in this model was the inclusion
Figure 7·19. Lead deposition as modeled by the updated (1996) hemispher-
of a particle size spectrum to the parameterization of the
ic EMEP transport model in a simulation for the reference year 1988.
The gridded Pb emissions data used in the model simulation are shown in
model. Particle size greatly affects the removal of heavy
Figure 7·5. (Source of data: Norwegian Meteorological Institute, Leonor
metals from the atmosphere. This model was then applied
Tarrasón pers. comm.).
386
AMAP Assessment Report
V
Mn
Zn
deposition process efficient enough to retain small particles
with heavy metals within the Arctic region, or are they car-
ried out of the region with air masses? Although the answer
to this question has is of fundamental importance when as-
sessing the impact of industrial emissions on the quality of
the environment in the Arctic region, there are only a few
measurements of wet deposition of heavy metals and other
pollutants, and even fewer of dry deposition. Generally,
As
In
Sb
these processes in the Arctic are poorly understood.
In general, models of long-range transport, deposition,
and modification of heavy metals in the Arctic are incom-
plete at the present. Bowling and Shaw (1992), for example,
have indicated that when physically reasonable constraints
are placed on sources and sinks of heat and water mass, and
on the relative humidity of near-surface air, it becomes ap-
parent that the assumption of isentropic flow without pre-
cipitation is incompatible with observed water mixing ratios
in layers of particles in the Arctic.
Al
Si
Sc
7.3.1.2. Receptor models to study the impact
of sources outside the Arctic
A number of statistical source-receptor techniques have been
used to study the contribution of various sources or even
source regions to the contamination of the Arctic air by heavy
metals. Maenhaut et al. (1989) applied the absolute princi-
pal component analysis (APCA) and the chemical mass bal-
ance method (CMB) to assess this contribution. The APCA
indicated that there were three source components contri-
Na
K
Ca
buting to the chemical composition of measured aerosols:
1) anthropogenic activity, 2) soil dust, and 3) sea-salt. The
average contributions of these components to the atmos-
pheric concentrations of several heavy metals in winter aero-
sols measured in the Norwegian Arctic is given in Figure
7·20 (after Maenhaut et al. 1989). The APCA does not,
however, produce a fine resolution of the contributions of
emissions from various source regions. To attempt this, the
CMB analysis of source apportionment was performed using
a set of elemental signatures obtained by Lowenthal and
Anthropogenic
Crustal
Sea-salt
Rahn (1985). Average regional apportionments of metals
from Europe and the Asian part of Russia to aerosols in the
Figure 7·20. Average contributions from various sources to the atmos-
pheric concentrations of elements in winter air samples at Ny-Ålesund,
Norwegian Arctic are shown in Figure 7·21 (after Maenhaut
Svalbard (after Maenhaut et al. 1989). The source contributions are ex-
et al. 1989).
pressed as the percentage of the average air concentrations; these did not
The North American path has been described in source-
always sum to 100% due to inaccuracy of methods.
receptor studies by Maenhaut et al. (1989). There are also
ter, 60 to 70% of heavy metal input is transported to the re-
gion from sources in the former Soviet Union, with the rest
100
from Europe and North America (Raatz 1984, Lowenthal
90
80
and Rahn 1985, Maenhaut et al. 1989). In summer, the con-
70
tribution from sources in Europe can be as high as 75%. Up
60
to 6% of the total emissions of As, Cd, Pb, Zn, V, and Sb in
50
all of Eurasia is deposited in the Arctic (Akeredolu et al. 1994).
40
It is important to note that the meteorological conditions
30
in the High Arctic during winter do not favor the deposition
20
10
of pollutants. Relative to other global locations, there are
0
fewer and smaller cloud droplets or ice crystals for particles
As
Pb
Zn
non-crustal
Se
non-crustal
to collide with or to diffuse and attach to in the polar air.
V
Mn
The Arctic air mass has great dynamic stability and is char-
Sources:
acterized by laminar flow. Washout of pollution is low
Eastern Europe
throughout most of the polar air mass. Stable stratification
in winter prevents strong vertical mixing (Shaw 1986, Ben-
Western Europe
Ny-Ålesund
son 1986). Under such conditions, pollutants (including
United Kingdom
heavy metals) transported to the Arctic can be trapped for
Central former Soviet Union
several weeks. Indeed, the episode of `mega-haze' in the Arc-
tic in the late winter of 1986 seems to confirm this hypothe-
Figure 7·21. Average regional source apportionment of metals in winter
aerosol at Ny-Ålesund, Svalbard from Europe and the central part of the
sis (Li and Winchester 1989). On some occasions, however,
former Soviet Union (including the Urals and Norilsk). (After Maenhaut
the episodes persist no longer than a few hours. Thus, is the
et al. 1989).
Chapter 7 · Heavy Metals
387
flows of air masses from the northern Pacific to Alaska
the case of Pyasino Lake, where heavy metals discharged with
(Shaw 1991b). This air system contains mean concentra-
waste waters from the Norilsk smelter tend to accumulate.
tions of heavy metals two to three times lower than the air
For pollution sources situated in the upper reaches of
mass system associated with lobes of the polar air mass that
large Arctic rivers (e.g. the Urals industrial complex in the
influence the area of central Alaska.
upper reaches of the Ob River basin) the transport of parti-
Information on the origin of heavy metals and other pol-
cle-bound heavy metals to the Arctic is either reduced or
lutants in the Arctic as described above has been confirmed
eliminated by deposition occurring along the river bed.
through the application of potential source contribution
Rivers are the largest source of freshwater to the Arctic.
function (PSCF) to the sets of data from the Norwegian Arc-
North-flowing rivers drain an area of 10 000 000 km2 of
tic (Li et al. 1996).
northern Asia, northern Europe, and North America as far
Other methods have been applied to study the origin of
south as latitude 50°N (Barrie et al. 1992). Most major
pollution in Arctic air and include:
north-flowing rivers are found in Asia and Europe with only
one, the Mackenzie, in North America. Cattle (1985) esti-
1. A tracer system based on ratios between air concentra-
mated that the total annual river discharge into the central
tions of various trace elements (Lowenthal and Rahn
Arctic Basin ranges from 2700 to 5000 km3 and varies from
1985);
year to year. Three Russian rivers (the Yenisey, Lena, and
2. Factor analysis techniques (Barrie and Hoff 1985);
Ob) account for almost 70% of the total water input to the
3. Individual particle analysis (Anderson et al. 1992, Sheri-
Arctic Ocean. Very few mass balance studies between the
dan 1989); and
different pathways are available. The relative importance of
4. Analyzes using information on isotope ratios in the Arctic
the different pathways depends on the metals considered
air (Sturges and Barrie 1989), halogens (Sturges and Bar-
and on whether figures for emissions or actual deposited val-
rie 1988), graphitic carbon (Rosen et al. 1981, Heint-
ues are used. In addition, the season and the distance from
zenberg 1982), and organic compounds (Pacyna and
the sources will strongly influence the estimates. A compari-
Oehme 1988).
son of the river fluxes of metals to the Arctic Ocean, pre-
These studies have proved that the Eurasian source regions
sented in Table 7·18, with atmospheric emission data from
contribute most of the contamination measured in the Arc-
Akeredolu et al. (1994) shows great variation in the relative
tic, particularly for heavy metals.
significance of these pathways. For metals such as Cd and
Pb, riverine transport toward the Arctic is approximately
half as great as the atmospheric contribution (Cd: 47 tonnes
7.3.1.3. Modeling the dispersion of emissions
in atmosphere, 25.6 tonnes in river; Pb: 2400 tonnes in at-
from sources in the Arctic
mosphere, 859 tonnes in river). Other metals show the re-
A Gaussian model has also been used to study the local im-
verse: for example, rivers carry five times more Zn than at-
pact of emissions from sources within the Arctic. Baklanov
mospheric emissions (1350 tonnes in atmosphere, 6660 ton-
and Rodjushkina (1993) studied the impact of emissions
nes in river).
from the Severonikel smelter complex on the Kola Peninsula
Although the amount of freshwater flowing into the Arc-
on the surrounding environment. The modeled average con-
tic Ocean equals only about 1% of the water masses enter-
centrations of Cu and Ni were higher than the measured
ing through Fram Strait, rivers have an important effect on
values by factors of 2 and 8, respectively. This rather high
the oceanography of the region. In draining vast areas of
disagreement can be partly explained by overestimation of
land, they may also carry significant amounts of contami-
emissions. As mentioned earlier, significant variability exists
nants to the ocean surface layer.
among emission estimates for this region.
The flow of freshwater to the Arctic is strongly seasonal.
For the Ob and Yenisey, the effects of spring thaw are evi-
dent in May, with a sudden increase to maximum flow in
7.3.2. Rivers
June. The relatively high monthly flow rates and lower but
As noted in chapter 3, rivers are one of the major pathways
more sustained flow of the Mackenzie, peaking in July, re-
of contaminants 1) to the Arctic region from the lower lati-
flect the moderating effects of the large lakes in the drainage
tudes and 2) to the Arctic marine ecosystem from the Arctic
system.
terrestrial ecosystem. The speciation of heavy metals in
rivers is controlled by the physico-chemical properties of
7.3.3. Estuaries
individual heavy metals and by river chemical conditions,
particularly the presence of complexing ligands such as
As noted in chapter 3, estuarine zones function as filters for
natural organic compounds. Concentrations of natural
terrigenous matter entering the ocean (marginal filter) and
organic matter in the Mackenzie River and in Siberian
combine the characteristics of both the catchment areas and
rivers are similar but much lower than those of the Rus-
the receiving marine basin. To assess a net flux of trace ele-
sian Euroarctic rivers (North Dvina and Pechora). Though
ments to the open ocean, the processes that control their
water discharges by the Siberian rivers are much higher
transport through an estuary must be understood.
than those of the North American rivers, sediment trans-
The Ob, Yenisey, and Lena estuaries are salt wedge estu-
port by Siberian rivers is significantly lower (except the
aries with minimal (generally < 1 m) tidal range, compara-
Yana, Indigirka, and Kolyma rivers in Eastern Siberia).
tively short freshwater residence-time, and high seasonal
For example, the Mackenzie River sediment load is seven
variability and biological productivity. Such estuaries pro-
times higher than that of the Yenisey River.
vide favorable conditions for highly variable chemical and
Sedimentation processes play an important role in the
biochemical reactions which strongly influence the behavior
fate of heavy metals in river transport, particularly if the
of many elements. Metals also exhibit individual tendencies.
rivers contain reservoirs or pass through lakes (e.g., Mac-
This is illustrated by the distribution of concentrations of
kenzie River-Great Slave Lake, Pyasina River-Pyasino Lake).
dissolved Cd, Cu, Ni, Pb, and Zn in the Lena mixing zone,
In this case, lakes provide a sink for heavy metals trans-
based on data obtained during a September 1989 expedition
ported by river flow. This situation is especially important in
(Martin et al. 1993). Cu, Ni, and Zn concentrations changed
388
AMAP Assessment Report
insignificantly with salinity. Lead concentrations showed
heavy metals released from discharge sources in Siberia.
strong salinity related variations and a tendency to be re-
Within the Arctic Archipelago, the net transport is from the
moved from solution, whereas Cd concentrations increased
Arctic Ocean to Baffin Bay.
with increasing salinity.
More detailed data on the fate of dissolved metals in the
7.3.5. Ice
Lena and Yenisey estuaries, which confirmed the conclu-
sions of the previous studies, were obtained during an inter-
Various pollutants, including heavy metals, deposited on the
national expedition to the Kara Sea in autumn 1993 (Lisit-
sea ice from the atmosphere can be transported long dis-
sin and Vinogradov 1994, Kravtsov et al. 1994, Dai and
tances in the ice or in blowing snow and then released to the
Martin 1995). In these studies it was shown that colloids
atmosphere or the ocean during the melt processes. In addi-
controlled the fate of dissolved metals in these estuaries as
tion, there is also a vertical transport, because growing sea
well as their net river flux to the open part of the Kara Sea.
ice rejects salt from the ice matrix in the form of dense brine
The behavior of Fe (removal) and Cd (desorption) was
which drains into the surface waters beneath. This vertical
consistent for the estuarine zones of all three largest Sibe-
convection thus enhances exchange between the surface and
rian rivers. Behavior of the other metals was specific for
deep ocean compartments (Gade et al. 1974).
each estuary. Changes of metal speciation in estuaries drasti-
Spring breakup of ice becomes evident from late April to
cally affects their fluxes in dissolved and suspended forms.
mid-May as flaw leads and polynyas expand. Under the in-
For Fe, Pb, and Zn for example, only 10-20% of total river
fluence of river runoff, landfast ice begins to deteriorate in
flux reaches the open ocean. In the case of Cu and Ni, the
June. In late summer, sea ice forms as the surface water cools
net flux is approximately the same as the total river input.
to its freezing point.
The flux of dissolved Cd to the sea tends to increase due to
According to Russian observations, Pb, Fe, Cu, and Cd
the desorption of Cd from suspended particles.
typically are elevated in sea ice/snow compared with surface
ocean water (Melnikov 1991), perhaps due to atmospheric
deposition and infreezing of particulate matter. Under-ice
7.3.4. Oceans
observations in the Laptev Sea indicate that starting in March,
Contaminants, including heavy metals in water or ice in the
Pb, Fe, and Cu are released from the ice, apparently due to
Arctic marine environment, are transported directly by ocean
brine migration, resulting in concentrations in the surface
currents. Ocean circulation is driven by a combination of
water that are two to three times higher than the initial val-
forces. A particular force can dominate in a particular geo-
ues (Pb increased from 0.1 to > 0.2 g/kg, Fe from < 0.5 to
graphical area; for example, tidal forces are dominant in ma-
>1.0 g/kg, and Cu from < 0.1 to 0.2 g/kg) (Melnikov 1991).
ny channels of the Arctic Archipelago whereas wind stress is
Campbell and Yeats (1982) concluded that, during melt-
most important for surface currents in the Canadian Basin
ing, ice contributed Fe, Cu, and Cd to surface waters in
(Barrie et al. 1992).
northwest Baffin Bay. In their study, sea ice with notably
The major water exchange between the Arctic Ocean and
high particulate concentrations (4.75 mg/L) collected off
other oceans occurs through Fram Strait. The West Spitsber-
Bylot Island yielded concentrations of these metals signifi-
gen Current flows northward off the west coast of Spitsber-
cantly in excess of the levels observed in surface waters, as
gen, transporting Atlantic water from the Nordic seas into
shown in Table 7·3.
the Arctic Ocean (Swift and Aagaard 1981). Relative to Arc-
Table 7·3. Concentrations of metals in ice and water, Baffin Bay (Campbell
tic Ocean water, the inflowing Atlantic water can be en-
and Yeats 1982).
riched with heavy metals, such as Pb, Cu, Ni, Hg, As, V, Cd,
Concentrations, in ice and water, µg/kg
and Cr, as seen from measurements carried out within the
Fe
Ni
Cu
Cd
Hg
Oslo and Paris (OSPAR) Convention on the reduction of the
contamination of the North Sea (ATMOS 1993).
Ice
1.32-1.47 25.28-59.90
0.37
8.22-7.29
0.31
Approximately half of the West Spitsbergen Current
Water
0.10-0.56
0.60-3.07
0.17-0.28
0.18-0.60 0.020-0.075
flows across the strait and recirculates southward at the
surface as modified (cooled) Atlantic water. The remainder
The polar ice moves continuously, under the influence of
flows northward, divides into various branches, cools on
the wind and ocean currents. The polar pack in the Beaufort
contact with the polar ice pack, and subsides beneath the
Sea drifts in a clockwise direction, whereas in the rest of the
Arctic surface water forming the deep ocean layer. Beneath
Arctic Ocean the transpolar drift carries ice from the Siberi-
the Western Spitsbergen Current lies northward-flowing
an coast toward Fram Strait where it exits the Arctic Ocean
bottom water. Together, these currents make up to 78% of
(Gordienko 1958).
the inflow to the Arctic Basin. On the western side of the
Fram Strait, the East Greenland Current flows southward at
the surface carrying sea ice to the North Atlantic. Beneath
7.4. Toxicological characteristics
the East Greenland Current are the southward-flowing modi-
7.4.1. Toxicokinetics: general principles
fied Atlantic water and bottom water. Together, these cur-
rents constitute about 83% of the outflow from the Polar
Heavy metals occur naturally in the Earth's crust and are
Basin (Barrie et al. 1992). Exchange with the Pacific Ocean
ubiquitous in the environment. Some heavy metals (e.g., Zn,
across the Bering Strait is one-way into the Arctic, contri-
Cu, Se) are biologically essential, playing an integral role in
buting about 20% of the total inflow. Again, large temporal
enzyme structure or as catalytic co-factors. Over 50 enzymes
and spatial variabilities occur in this flow (Coachman and
require Zn as a co-factor for example, and no life can exist
Aagaard 1981).
without it. Other metals, such as Pb, Cd, and Hg, do not ap-
Within the Arctic Ocean itself, the main surface circula-
pear to have any biological function in organisms and are
tion features are the clockwise circulation of the Beaufort
accordingly termed non-essential metals. Only trace amounts
Gyre and the transpolar drift. The transpolar drift flows
of essential heavy metals are required physiologically; in in-
from Siberia, across the pole, then southward to exit as the
creased doses, these metals become toxic, as do non-essen-
East Greenland Current. This drift can be contaminated by
tial metals. Animals have developed a variety of homeostatic
Chapter 7 · Heavy Metals
389
mechanisms for the regulation of levels of essential metals
other complexing ligands, suspended particulate matter, tem-
(Clarkson 1986), which usually involve the control of gas-
perature) influencing the concentration of divalent cations in
trointestinal absorption. Toxic effects are therefore less
solution will affect bioavailability. In animals, the major
likely to be experienced with these metals than is the case
routes of heavy metal uptake in order of efficiency are inges-
with non-essential heavy metals.
tion, inhalation, and dermal absorption (see Table 7·5).
The three non-essential heavy metals that have been iden-
Heavy metals tend to accumulate in so-called `storage'
tified as being of greatest concern in the Arctic, because of
compartments as long as the rate of uptake exceeds the rate
their consequences for human health (see chapter 12) are
of excretion. Because rates of excretion or elimination of
Pb, Cd, and Hg. Selenium has also been included for discus-
metals are often different in different parts of the body, these
sion because it is essential in enzymes and proteins, and can
storage compartments represent the lumping together of var-
reduce the toxic effects of some heavy metals (e.g., Cd, Hg,
ious parts of the body where rates of specific metal elimina-
and As see below). Similarly, the well-defined toxic effects
tion are the same. Accumulation often continues throughout
of Se can be reduced by some of these metals. In contrast to
life, or in some cases, at least until metal toxicity is experi-
most organic chemicals (see chapter 6) that can be elimi-
enced. Pb, for example, is typically redistributed from var-
nated from body tissues via metabolic degradation, elemen-
ious organs into two tissue compartments: an exchangeable
tal metals cannot be degraded. The only mechanism by
compartment (blood and soft tissue) and a storage compart-
which metals can be removed from the body is by excretion.
ment (bone). While accumulation occurs in both compart-
Heavy metals, therefore, have the potential to accumulate in
ments, Pb levels tend to stabilize at a maximum level in the
the body, leading to acute and chronic effects.
exchangeable compartment with time, or in some cases even
The high reactivity of heavy metals has a direct bearing
to decrease. Lead levels in bone, however, tend to increase
on their bioavailability, distribution, and ultimate toxicity.
throughout the lifetime of an organism.
The major factor underlying the biochemical properties of
heavy metals with regard to their transportation, distribu-
7.4.2.1. Bioaccumulation and biomagnification:
tion, and elimination in organisms is their high affinity to
general principles
sulfur and sulfhydryl groups of proteins. Sulfhydryl groups
are ubiquitous in organisms, occurring on plasma proteins,
The ability of organisms to accumulate heavy metals to con-
membrane proteins, and enzymes.
centrations of one or more orders of magnitude greater than
Factors influencing the toxicity of heavy metals in solu-
concentrations in their food usually represents the major
tion are summarized in the following scheme (after Bryan
pathway leading to chronic toxicity. However, accumulated
1976):
metal may be present in tissues in a relatively non-toxic or
inert form even if it was originally toxic, because the toxicity
Ion
of the metal can be modified through interactions between
Soluble
Complex ion
metals or through biotransformation by the organism. In
Chelate ion
contrast to POPs which are highly lipophilic and therefore
Form of metal
Inorganic
Molecule
in water
Organic
accumulate primarily in body lipids, heavy metals are prefer-
Colloidal
entially accumulated in proteinaceous tissues. The degree to
Particulate
Precipitated
which metals are accumulated varies greatly depending both
Adsorbed
on the metals involved and on the organ or tissue. The con-
Presence of
Joint action
More-than additive
cepts of bioaccumulation and biomagnification used in stud-
other metals
No interaction
Additive
ies of the environmental distribution of heavy metals are
or poisons
Antagonism
Less-than-additive
slightly different than those used for other persistent pollu-
tants and therefore need to be clarified to avoid confusion.
Factors influencing
Temperature
The term bioaccumulation refers to the net accumulation
physiology of
Ph
organisms and
Dissolved oxygen
of metals within an organism from both biotic (other organ-
possibly form of
Light
isms) and abiotic (soil, air, and water) sources. Bioaccumu-
metal in water
Salinity
lation (or the bioaccumulation factor, BAF) is expressed as
the relationship between the concentration of a metal in the
Stage in life history (egg, larva, etc.)
Changes in life cycle (e.g. molting, reproduction)
tissue of an organism and the concentration of the metal in
Age and size
the air (terrestrial) or water (aquatic). This unitless ratio rep-
Condition
Sex
resents the steady state between the rate of uptake and the
of organism
Starvation
Activity
rate of loss of metal by the organism, and is largely depen-
Additional protection (e.g. shell)
dent on the bioavailability of different metals to an organ-
Adaptation to metals
ism, and the tissue type selected for measurement. Bioaccu-
Behavioral response
Altered behavior
mulation can vary naturally between individual organisms
and within populations just as metal concentrations can
vary among individuals of the same species; it depends part-
ly on an organism's food habits and overall physiological
7.4.2. Uptake
condition.
Uptake of heavy metals is directly related to bioavailability.
Generally speaking, the BAF is more appropriately used
Factors affecting bioavailability include both those specific
to describe accumulation in the aquatic environment, as the
to the organism (age, feeding habits, and nutritional status)
uptake of heavy metals by higher organisms in the terrestrial
and those specific to the metal (chemical form, particle size)
environment occurs primarily via ingestion. Heavy metal con-
(Beck et al. 1995). In general, the uptake of metals by biota
centrations in the surrounding terrestrial environment (soil
involves crossing a biological membrane. This is most read-
and air) play a less important role in bioaccumulation than
ily achieved by divalent cations. Consequently, any of the
is the case in water. A concept closely related to bioaccumu-
physico-chemical variables (e.g., pH, redox potential, ionic
lation, `bioconcentration' (or bioconcentration factor, BCF)
strength, presence/absence of dissolved organic matter and
describes the concentration of a metal, derived only from
390
AMAP Assessment Report
water (through gills or epithelial tissue), in a tissue relative
Terrestrial lichens, which lack root systems, take up met-
to the concentration of the metal in the water (Macek et al.
als directly from the air. While absorbing a small proportion
1979). Bioconcentration can, realistically, only be measured
of heavy metals through their leaves, vascular plants gener-
in the laboratory under specific conditions. Although dermal
ally absorb metal cations through their roots from soil wa-
absorption may play an important role in aquatic systems,
ter. Fungi accumulate metals differently than vascular plants,
most organisms (except phytoplankton) do not necessarily
showing some degree of selectivity. Most macrofungi species
accumulate metals in tissues directly and solely from water.
contain significantly more Zn and Cd than plants, and cer-
The progressive bioaccumulation of heavy metals by suc-
tain species strongly accumulate Hg and Cd (Lodenius 1990).
cessive trophic levels is termed biomagnification. The bio-
Direct uptake of metals from the air through the lungs is of
magnification factor (BMF) is generally expressed as the
no significance in the Arctic environment, except in the im-
ratio of two heavy metal concentrations (Cn and Cn+1) pre-
mediate vicinity of major pollution sources.
sent in the tissue at two trophic levels (n and n + 1). In par-
ticular, it can be described as the concentration ratio in a tis-
7.4.2.3. Freshwater ecosystem:
sue of a predator organism to that in a prey organism (i.e.,
bioaccumulation/biomagnification
Cn+1/Cn). The BMF is not restricted to contaminant move-
ment over a single trophic level, and can in fact refer to con-
The main sources of heavy metal contamination in the fresh-
centrations in organisms one or more trophic levels apart.
water environment include atmospheric fallout, smelting pro-
Although heavy metals have been found in most tissues
cesses, mining activities, and runoff from terrestrial drainage
of mammals, certain metals tend to accumulate preferen-
systems. Aquatic systems differ from terrestrial ones in the
tially in one or more tissues. For instance, in most mammals
proportion of metals they receive from natural sources (back-
experiencing chronic exposure to Cd, concentrations occur
ground concentration) versus anthropogenic sources.
primarily in the kidney. Only small, relatively insignificant
The acidification of terrestrial soils by acid rain increases
concentrations are present in muscle. Similarly, Hg is accu-
the solubility of certain metal ions including Mn, Zn, Pb,
mulated largely in the liver, and Pb in bone. It is therefore
and Cd, thereby increasing leaching into surface waters. In a
important to compare metal concentrations from the same
study affecting heavy metal bioaccumulation in pristine
tissue compartment for both predator and prey in order to
headwater lakes, Iivonen (1992) concluded that the bioaccu-
obtain consistent BMFs. Inconsistencies can arise in the
mulation of several metals, particularly Pb and Cd, was in-
comparison of biomagnification factors due to the selection
creased with increasing acidification of the water column.
of different tissues for contaminant measurements. The
Furthermore, there was some evidence that the presence of
BMF for Cd in polar bears (Ursus maritimus) relative to
organic matter reduced heavy metal bioaccumulation, prob-
seals is < 1 if based on liver tissue of both animals, but is 77
ably due to complexation of heavy metals with dissolved
if based on blubber, the tissue (along with skin) preferen-
humic matter. The stabilities of metal-humic complexes in
tially consumed by bears (Muir et al. 1992).
natural waters are higher than those of most corresponding
inorganic metal complexes.
Heavy metal uptake by aquatic biota begins with the low-
7.4.2.2. Terrestrial ecosystem:
er trophic levels. Unicellular organisms (bacteria, algae) and
bioaccumulation/biomagnification
aquatic plants take up metals exclusively from the water
Background levels of heavy metals in the soil vary region-
(pore water in the case of rooted plants) by passive surface
ally, and are primarily affected by the geology of local bed-
absorption. Organisms at higher trophic levels generally ac-
rock. The anthropogenic contribution to heavy metal conta-
cumulate metals via both passive absorption directly from
mination through long range atmospheric fallout (from pre-
the water and via food items. Zoobenthic and zooplanktonic
cipitation and dry deposition) exceeds the natural compo-
organisms take up metals largely by feeding on algae, partic-
nent (from degassing of the earth's crust) (Pacyna 1994). In
ulate organic bottom and suspended matter, and on other
the 1950s and 1960s, the use of heavy metals in agricultural
zoobenthic and zooplanktonic organisms. The extent of
chemicals (the use of alkylmercury fungicides as seed dress-
heavy metal uptake over body surfaces is not known, but is
ings in particular) resulted in intoxication of the terrestrial
likely related to metal concentration and chemical speciation
food chain. Industrial sources of contamination (discussed in
in the water column.
section 7.6.2.1), so-called `hot spots,' are generally localized;
Carnivorous fish, at the highest trophic level in the fresh-
deposition rates are markedly elevated nearest the source
water ecosystem, absorb metals from the water partly through
and decrease rapidly with distance.
body surfaces, primarily the gills, and partly from food (zoo-
Most heavy metals deposited in soil-water systems are
plankton, zoobenthos, and smaller fish). The proportion of
rapidly bound by organic particulate matter. The avail-
uptake attributable to either source varies according to the
ability of these metals to terrestrial microorganisms depends
species of fish, the metal, and the concentrations in the wa-
largely on the their oxidation state, the organic content of
ter and in the food. In one species of fish, 90% of the inor-
the soil, and pH. In the Arctic, soil acidification due to acid
ganic Hg burden was absorbed directly from water (Jerne-
rain is an important factor contributing to heavy metal bio-
low 1972), whereas in another species, more than 80% of
accumulation, as most metals are more readily accumulated
MeHg was absorbed from food (Hall et al. 1997).
as divalent cations, and acidification generally enhances the
concentration of divalent cations in solution. Mercury be-
7.4.2.4. Marine ecosystem:
haves in the opposite way: decreasing pH enhances the ab-
bioaccumulation/biomagnification
sorption of Hg onto organic matter (Bergkvist 1986, Lode-
nius 1990). At pH 4.8 Bergholm et al. (1985) found the fol-
Heavy metals enter the marine environment through various
lowing binding strength sequence based on the solubility of
pathways including atmospheric deposition, transport by
metals: Fe > Al > Cu > Pb > Zn > Mn = Cd. Similarly, An-
rivers, leaching from agricultural land, weathering of rocks
dersson et al. (1991) suggested the following sequence in the
and soils, and from point sources (e.g., mining, spills, incin-
pH range 3.5-5.0: Hg > Pb > Cu > Zn > Cd > Mn > Ca > Mg,
eration of wastes on the open sea, and various accidental or
with Hg being bound most strongly.
deliberate discharges). Studies monitoring heavy metal con-
Chapter 7 · Heavy Metals
391
tamination in European seas indicate that atmospheric path-
Table 7·4. Biomagnification factors of cadmium, mercury, and selenium in
ways (atmospheric particles, gas exchange) play a dominant
the Greenlandic marine food chain.
role in the transport of Pb, Cd, As, Cu, Zn, and other ele-
Biomagnification factor,
ments. In the North Sea and the Baltic Sea approximately
wet weight basis
50% of these elements is deposited from the atmosphere
Species/tissue
Area
Cd
Hg
Se
(ATMOS 1993, Pacyna 1992). Inputs of metals to coastal
Parathemisto vs. Arctic cod
waters via surface runoff from north-flowing rivers are im-
Whole vs. liver
Avanersuaq
0.9
6.4
3.8
portant contributors to the sea in the summer, where they
Whole vs. muscle
Avanersuaq
0.04
15.6
2.3
accumulate in sediments and in marine biota.
Blue mussel vs. common eider a
The food webs of the Arctic depend mainly on primary
Soft tissue vs. liver
Qeqertarsuaq
5.1
49.8
5.8
production by planktonic organisms and ice algae. This pro-
Nanortalik
4.1
47.6
6.7
Soft tissue vs. kidney
Qeqertarsuaq
12.7
16.6
6.0
duction is highly seasonal and depends on local conditions
Nanortalik
17.1
18.4
5.6
(such as temperature regime and nutrient enrichment). Phy-
Soft tissue vs. muscle
Qeqertarsuaq
0.07
8.1
1.0
toplankton blooms assimilate heavy metals from the water
Nanortalik
0.13
10.1
0.7
column and either transfer metals directly to higher trophic
Arctic cod vs. black guillemot
Liver vs. liver
Avanersuaq
4.1
46
3.0
levels via grazing (e.g., zooplankton), or transport them to
Ittoqqortoormiit
9.8
63
2.7
the ocean floor via processes of bodily absorption and sink-
Muscle vs. muscle
Avanersuaq
6.1
5.6
1.7
ing. Similarly, as upwelled water ages, organisms gradually
Ittoqqortoormiit
22
6.8
2.9
enrich the water with organic compounds, some of which
Arctic cod vs. Brünnichs guillemot
Liver vs. liver
Avanersuaq
7.9
56
2.3
act as chelating agents binding to metals, thereafter trans-
Ittoqqortoormiit
24.6
118
3.1
porting them to the sea floor as particulate matter. In this
Muscle vs. muscle
Avanersuaq
11
5.9
1.5
way most of the sediments in areas of upwelling are gener-
Arctic cod vs. northern fulmar
ally rich in metals.
Liver vs. liver
Avanersuaq
16.3
112
10.7
Once on the bottom, metals may remain at the surface
Ittoqqortoormiit
18.9
170
5.9
Muscle vs. muscle
Avanersuaq
20.6
8.1
5.0
as a result of biological mixing (bioturbation) of the sedi-
Ittoqqortoormiit
64.4
10.2
12.4
ments by benthic invertebrates, which in turn support bot-
Arctic cod vs. ringed seal b
tom feeding animals such as bearded seal (Erignathus barba-
Liver vs. liver
Avanersuaq
19.9
115
1.8
tus), walrus (Odobenus rosmarus), and some whale species.
Nanortalik
20.8
636
3.5
Kong O. Fjord
6.8
2823
10.9
Consequently, metals may remain in contact with benthic bio-
Ittoqqortoormiit
30.6
840
2.7
ta for longer periods of time than if buried. Dead detrital or-
Arctic cod vs. narwhal c
ganic particles are mineralized and returned to the water col-
Liver vs. liver
Avanersuaq
13.3
178
3.5
umn where they again become available for bioaccumulation.
Muscle vs. muscle
Avanersuaq
1.9
6.6
0.4
The High Arctic food web can have as many as five steps,
Ringed seal vs. polar bear d
from algae through invertebrates to fishes, seals, and finally,
Liver vs. liver
Avanersuaq
0.13
8.9
3.6
Ittoqqortoormiit
0.06
1.8
1.3
polar bears (Welch et al. 1992). Arctic cod (Arctogadus gla-
Kidney vs. kidney
Avanersuaq
0.29
14.4
3.3
cialis) is a pivotal species in the food web, and the polar
Ittoqqortoormiit
0.28
9.2
1.3
bear is the apex predator.
Muscle vs. muscle
Avanersuaq
0.13
0.28
1.8
Ittoqqortoormiit
0.20
0.24
0.85
In Table 7·4, BMFs for some key species in the Green-
landic marine ecosystem have been listed. At the lower tro-
a. Means of two size groups (30-40 and 40-50 mm) of blue mussels from
Nanortalik and only one size group (40-50 mm) from Qeqertarsuaq
phic end, it is obvious that Cd is not accumulated from pa-
were used for calculations.
rathemisto to Arctic cod, even when comparing liver of Arc-
b. Means of 4-5 age groups of ringed seals were used for calculations.
tic cod with the whole parathemisto. On the other hand, a
c. Means of juvenile, mature and adult narwhals were used for calculations.
d. Means of 4-5 age groups of seals and 3 age groups of polar bear were
clear accumulation takes place in birds, seals, and whales,
used for calculations.
where the BMF may be as high as 64.4. for muscle in fulmar
(Fulmarus glacialis) versus Arctic cod. The BMF is also low
by various organs and the fraction subsequently excreted
between polar bears and ringed seals (Phoca hispida) be-
vary greatly for each metal. The mechanisms involved in
cause polar bears prefer to eat seal blubber, which is low in
selective metal uptake in organs are not well understood
Cd (Muir et al. 1992, Dietz et al. 1995). Muir et al. (1992)
(Foulkes 1995). Metals are initially distributed to a variety
also reported similar BMF for Parathemisto libellula/Arctic
of organs and tissues, and subsequently redistributed to
cod and Arctic cod/narwhal (Monodon monoceros) from the
other tissues for storage and inactivation. For example, Cd
Canadian High Arctic.
in the blood (bound to high molecular weight proteins) ini-
Mercury increases from 6.4 to 15.6 times in liver and mus-
tially accumulates in the liver (where efficient metallothio-
cle of Arctic cod relative to whole parathemisto. The bio-
nein synthesis probably takes place), and is subsequently
magnification is much higher between Arctic cod and ringed
redistributed to the kidney as a Cd-metallothionein com-
seal, where a BMF can reach 2823. In general the Se tends
plex. In the latter form, Cd can be readily filtered through
to biomagnify, but to a much lesser degree than does Hg.
renal glomerelli and reabsorbed in the tubuli where it binds
Insufficient data were available to produce similar figures
to the renal tubular cells (Friberg et al. 1986, Norberg and
for Pb from the same areas. However, data from Muir et al.
Norberg 1987). A similar mechanism is found with Cu,
(1992) and the data compiled in this assessment indicate no
which is initially bound to albumin and some amino acids
biomagnification, i.e., BMF < 1.
and transported in the plasma to liver, brain, and muscle tis-
sue. Under certain circumstances Cu can be remobilized by
7.4.3. Transport, biotransformation,
binding to a high molecular weight protein, ceruloplasmin,
and distribution
and released back into circulation to be transported to var-
ious other tissues.
Once absorbed, heavy metals are distributed in the body by
The formation or breakdown of metal-carbon bonds or a
the circulatory system, irrespective of their chemical form
change in the oxidation state of a metal within an organism
(Foulkes 1995). The fraction of transported metal absorbed
(biotransformation) will affect the chemical activity of heavy
392
AMAP Assessment Report
Table 7·5. Routes and efficiency of intake of Pb, Cd, Hg, and Se for different organisms. Information was compiled from reviews by WHO (1976, 1987,
1989a, 1989b, 1990, 1991, 1992a, 1992b), unless specified otherwise.
Metal
Organism Route
Compound
Efficiency
Reference
Pb
Terrestrial plants
Roots
Readily absorbed
Shoots
Lesser extent
Birds
Intestine
Dependent on fiber in diet
Animals
Intestine
Airborne Pb
5-10%
1
Humans
Lungs
Airborne Pb
30-50%
1
Intestine
10%
1
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Cd
Fish
Intestine
1%
Gills
0.1%
Rapidly exchanging tissue
Intestine
1-2%
2
Cows
Intestine
16%
2
Rats, mice, and monkeys
Intestine
1-6%
3
Animals
Lungs
7-40%
3
Humans
Lungs
25-50%
3
Intestine
5-7%
3
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Hg
Shrimp
Intestine
Methyl Hg
70-80%
Intestine
Inorganic Hg
38%
Fish
Gills
methyl Hg
High. at low temp. wat. hardn.
Intestine
methyl Hg
Higher at higher temp.
Birds
Intestine
Organic > inorganic
Mammals
Intestine
Methyl Hg
>95%
Intestine
Inorganic Hg
>15%
Humans
Lungs
Hg vapor
80%
4
Intestine
Metallic Hg
Poorly absorbed
4
Skin
Likely; no qual. data
4
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Se
Rats
Lungs
H2Se04
94%
4
Lungs
Elemental Se
57%
4
Intestine
Selenite
92%
Intestine
Selenomethionine 91%
Intestine
Selenocystine 81%
Intestine
Se in rabbit kidney
87%
Intestine
Se in fish muscle
64%
Skin
Selenite
10%
Sheep and cows
Intestine
30-35%
4
Mammals Intestine
Reg.
homeostatically
4
Humans
Intestine
Selenite
60%
Intestine
Selenate
94%
Intestine
Selenomethionine 96%
1. Tsuchiya (1986); 2. Paludan-Müller (1993); 3. Friberg et al. (1986); 4. Berlin (1986); 5. Högberg and Alexander (1986).
metal compounds, and therefore their toxicity. Changes in
as Cd, Zn, Cu, and Hg, for example, is fundamental to
the oxidation state influence the ability of a metal to interact
their toxicity. Metals binding with metallothionein form
with various tissue ligands. Hg, for example, exists in three
inert complexes which can be retained in body tissues
oxidation states: elemental (Hg(0)), the mercurous ion
(Clarkson 1986). Similarly, Se can reduce the toxicity of
(Hg2(II)), and the mercuric ion (Hg(II)). Hg(0) easily pene-
certain metals such as As, Cd, and Hg by forming inert
trates biological membranes because of its high lipid solubil-
compounds (Högberg and Alexander 1986) which usu-
ity. The mobility of Hg2(II) and Hg(II) are much more re-
ally accumulate within organisms.
stricted due to their tendency to form salts and their high
The fact that many Arctic animals at high trophic levels
affinity for sulfhydryl groups on proteins (Clarkson 1986).
(e.g., seals, whales, seabirds) are consumed by Arctic people
Two biotransformation processes are important to the
has provided the impetus for analyzing tissue concentrations
toxicity of heavy metals:
of various pollutants. Until now, measurements have been
1. Methylation/demethylation of certain heavy metals and
primarily carried out on those organs and tissues that are
metalloids (e.g., As, Hg, and Se). In some cases cleavage
consumed. As a result of, the literature on metals and other
may serve as a detoxification pathway, whereas in others
contaminants in the tissues of Arctic animals may or may
the metabolite is the more toxic species. For example,
not represent targets of toxicological action in the animals
methylation (the formation of metal-carbon bonds) of in-
themselves. Arctic fish and marine mammals accumulate rel-
organic As and Se has been seen to lead to reduced toxi-
atively high levels of methylmercury in kidney, liver, and
city in a number of animals and to form the basis for ex-
muscle. Although knowledge of these concentrations is im-
cretable metabolites (e.g., methylated selenides), while the
portant for assessing the potential exposure of human con-
reverse is true for Hg. In the case of methylmercury expo-
sumers to methylmercury, in terms of evaluating the toxic ef-
sure, processes of demethylation are important for detox-
fect to mammals the most important concentration is that in
ification. In the case of Se, biotransformation in the liver
the brain, which is generally not measured. A considerable
seems to be the major mechanism by which homeostasis
amount of data has been collected for the critical organ of
is maintained.
chronic Cd intoxication, the kidney. Even though Cd con-
2. Formation of inert complexes also plays an important
centrations exceeding the threshold for toxic effects in var-
role in heavy metal detoxification. The biochemical rela-
ious marine mammals have been measured, no clinical toxi-
tionship between metallothionein and such heavy metals
cological investigations have been carried out.
Chapter 7 · Heavy Metals
393
Table 7·6. Biological half-life of Pb, Cd, Hg, and Se for different organ-
bonds with non-absorbable compounds. In this way a net
isms and tissue groups. Information was compiled from reviews by WHO
gastrointestinal excretion can occur, with the heavy metals
(1976, 1987, 1989a, 1989b, 1990, 1991, 1992a, 1992b), unless specified
being eliminated from the body with feces. The other mech-
otherwise.
anism of gastrointestinal excretion involves the removal of
Metal Organism/tissue
Half-life (t 1/ )
Ref.
2
metal compounds in association with the rapid turnover of
cells of the intestinal mucosa. Small quantities of certain
Pb
Blue mussel
Trimethyl/ethyllead: 3-4 days
Insects:
heavy metals (e.g., inorganic Cd and Hg) are eliminated
Fast body burden
7 days
from the body when the cells are shed.
Slow body burden
22 days
Urinary excretion is probably the second most important
Fish tissue:
9 days (received from water)
40 days (received from diet)
excretory route for animals. The glomerula membrane acts
Trimethyl/ethyllead: > 41 days
as a filter, allowing only those molecules with relatively low
Humans:
molecular weights to pass through into the renal tubules.
Rapidly exchanging tissue 35 days
1
Soft tissue compartment
40 days
1
Thus, metals bound to low-molecular weight proteins such
Bone
20 years
1
as insulin or metallothionein may be cleared from the blood
-------------------------------------------------------------------------------------------------
Cd
Invertebrate tissue
2-53 days
plasma in this way, although a proportion of this is subject
Isopods and mollusks
Little or no excretion
to reabsorption. For example, Cd bound to metallothionein
Fish tissue
24-63 days
is efficiently reabsorbed, with only a small proportion being
Mice and rats
200-700 days
Squirrel monkeys
> 2 years
ultimately excreted with the urine.
Other mammals:
Liver
10-50% of lifespan
Slowest component
> 20% of lifespan
7.4.5. Uptake, accumulation, and loss in biota
Humans:
Kidney
10-30 years
2, 3
The following is a brief overview of the toxicokinetic char-
Liver
5-15 years
3
-------------------------------------------------------------------------------------------------
acteristics of the four heavy metals identified as most signifi-
Hg
Leguminous plants
Up to 75% as elemental Hg
cant to Arctic biota: Pb, Cd, Hg, and Se. The details pre-
Invertebrates
t1/2 inorganic > t1/2 organic
sented below are derived primarily from data contained in
Blue mussel
53-293 days
Fish:
Metallic and inorganic Hg:
the International Programme on Chemical Safety (IPCS) of
45 days (received in water)
the World Health Organization (WHO).
Organic Hg:
An important factor in the process of bioaccumulation
323 days (received from diet)
Inorganic Hg:
(which also has implications for toxicity) is the biological
61.6 days (received from diet)
half-time. This is the time it takes for the concentration of
Mice and rats
12 days
an absorbed pollutant to be reduced to half of its initial val-
Seals
Methyl Hg: 500 days
4
Dolphins
Methyl Hg: 1000 days
4
ue, and consequently biological half-times generally refer to
Humans:
rates of elimination from specific storage compartments.
Phase I: blood
Metallic Hg: 2-4 days
Rates of excretion vary between metals and their compounds,
(received as vapor)
Methyl Hg: 39-70 days
and may have a major influence on chronic toxicity. Most of
Phase II
Metallic Hg: 15-58 days
the data on the half-time of many heavy metals have been
(received as vapor)
Whole body
Methyl Hg: 52-93 days
obtained from actual measurements of elimination rates fol-
Inorganic Hg: 40 days
lowing exposure.
Kidney
Inorganic Hg: 64 days
A summary of the primary routes of uptake and biologi-
Brain
19-26 days
Hair
Bimodal:
cal half-times for Pb, Cd, Hg, and Se in various organisms
35-100 days and 110-120 days
are given in Tables 7·5 and 7·6, respectively. Invertebrates
-------------------------------------------------------------------------------------------------
Se
Rats:
and fish take up only a small proportion (0.1% for Cd) of
Kidney
38 days
the metals in the water through the gills. Intestinal intake by
Muscle
74 days
invertebrates and fish is larger; the most efficient intake is
Whole body
55 days
Animals:
for methylmercury (70-80%) and a much smaller uptake for
Phase I
1-3 days
5
Cd (1%). Excretion by invertebrates and fish compared with
Phase II
30-70 days
5
higher trophic levels is rather fast, with biological half-times
Humans:
Phase I
1 day
5
of 3-40, 2-63, and 53-323 days for Pb, Cd, and Hg, respec-
Phase II
8-20 days
5
tively. This means that the major route of metals is through
Phase III
65-116 days
5
food, where the levels also are higher than in the surround-
1. Carlson 1986 cited in WHO 1989a; 2. Chowdhury and Chandra 1987;
ing water.
3. Nordberg et al. 1985; 4. André et al. 1990; 5. Högberg and Alexander
For higher organisms, metals are efficiently (7-94%) ab-
1986.
sorbed through the lungs, with the degree of absorption var-
ying by animal species, metal, and chemical form of the met-
al. Mercury vapor and H
7.4.4. Excretion
2Se04 are taken up most efficiently
(80-94%), but Pb and Cd readily cross the lung epithelium
The most important excretory pathways for metal com-
(7-50%). Air levels of these metals, however, are rather low
pounds in animals are gastrointestinal and renal. Gastroin-
in most Arctic areas, so this pathway is of only minor im-
testinal excretion includes excretion of metals into bile (and
portance. The uptake through the intestine depends on the
pancreatic fluid) and excretion by the intestinal mucosa. Ex-
organism, metal type, and chemical form. Only 1-16% of Pb
cretion of most organic metal compounds occurs primarily
and Cd are taken up by various species, whereas 57-95% of
by the bile, while inorganic compounds are excreted in the
Hg and Se are absorbed through the intestine. Methylmer-
gastrointestinal tract. Considerable quantitative differences
cury is taken up more than six times as efficiently as inor-
have been reported for different animal species. Metals ex-
ganic Hg. Lead and Se are readily excreted, which results in
creted in the bile may be reabsorbed farther down the in-
moderate levels of these metals in internal organs. Lead,
testinal tract, and subsequently re-excreted into bile. Intesti-
however, is deposited in bone, where its half-time can be up
nal reabsorption can be prevented if excreted metals form
to 20 years. Cadmium is excreted slowly once it is taken up
394
AMAP Assessment Report
by organisms, with a half-time of 10-50% of the organism's
ity. Increased organic content of the water generally de-
life span (up to 30 years), whereas Hg has intermediate half-
creases uptake and toxicity. Complexes with EDTA, NTA,
time (12-1000 days). Differences in half-time between the
and DPTA, for example, render Cd unavailable, and com-
different species partly explain the generally higher observed
plexes with fulvic acids of low molecular weight and citric
metal levels in longer-lived species. The different half-times
acid make Cd less available than in its free ionic form. The
also explain differences between the observed levels of the
formation of hydrophobic complexes with compounds such
different metals in the various tissue compartments. One
as xanthates (used in mining for the enrichment of metals
might speculate that the excretion of Hg in seals and whales
from sulfide ores) and dithiocarbamates, on the other hand,
is so inefficient because of the abundance of Se in the envi-
enhance uptake by acting as carriers over biological mem-
ronment. Se detoxifies Hg by forming the biologically inert
branes, and are therefore of particular environmental con-
compound mercuric selenide, which is not destroyed by the
cern. In terrestrial systems, the major source of exposure to
proteolytic enzymes and is thus a terminal phase in the reac-
animals is via inhalation and ingestion.
tions of methylmercury detoxification in the liver (Martoja
Uptake in plants depends on the availability of ionic Cd,
and Berry 1980, cited in Andre et al. 1990).
which is affected by soil composition (i.e., particle size, min-
eralogy, organic matter content, and pH). Studies with lettuce
and chard reveal dose-related uptake from soil into leaves.
7.4.5.1. Lead
Red maple, white pine, and Norway spruce seedlings cul-
Lead in the environment is strongly absorbed onto sediment
tured in contaminated soil contained more Cd in the roots
and soil particles, and is therefore largely biologically un-
than in the leaves. Cadmium is usually found bound to the
available. Many of the inorganic salts formed (Pb oxides
cell wall in roots.
and sulfides) are not readily soluble in water and are se-
Most Cd present in mollusks is taken up directly from the
questered in sediments. In aquatic systems, uptake is influ-
water. Uptake in oysters depends largely on the rate of water
enced by various environmental factors such as temperature,
flow through the body, which is linked to feeding activity
salinity, pH, and the presence of organic matter. Among all
and water temperature. Accordingly, oysters can accumulate
the chemical species, Pb(II) is considered to be most readily
twice as much Cd in summer as in winter. Cadmium accu-
absorbed by biota.
mulates in both soft tissues and the shell, although localized
It has been shown that in terrestrial plants, Pb is taken up
accumulation can occur in the gills, heart, mantle, and mus-
through the roots and to a lesser extent through the shoots.
cles. Terrestrial pulmonate snails were found to absorb Cd
Although there has been some evidence of translocation, it
from the chloride administered in their diet. In the terrestrial
appears that this plays a very limited role in distribution.
slug (Arion ater) most of the Cd retained in the body was
Some of the Pb entering plant root cells becomes incorpo-
found associated with the digestive gland. Specific metallo-
rated in the cell wall. Studies of aquatic plants have shown
thionein-like Cd binding proteins have been isolated from
that submerged species often retain greater concentrations
both marine and terrestrial gastropods. Differences in the
of Pb than floating species, with both shoots and roots evi-
ability of marine crustaceans to absorb Cd directly from the
dently playing an important role in absorption.
water are attributed to variations in cuticle permeability and
For many animals it is not known for certain whether Pb
lifestyle. Accumulation is greatest in the hepatopancreas,
is absorbed through the skin or actually taken up via inhala-
with lower concentrations occurring in the exoskeleton,
tion or contaminated food. Accumulation in mussels (My-
muscle, and serum. There is no evidence that these organ-
tilus edulis) has been demonstrated to occur in all tissues,
isms possess any mechanisms for Cd regulation. Isopods and
but highest concentrations are seen in the kidney. In fish, Pb
mollusks retain Cd in their tissues with little or no excretion.
accumulates primarily in the gill, liver, and kidney, though it
Elimination half-lives in other species have been estimated to
is not known whether accumulation in the gills represents
be between 2 and 53 days.
uptake into the tissue or absorption onto exterior surfaces.
As with mollusks, metabolic rate and feeding are impor-
Birds dosed with lead shot show signs of tissue accumula-
tant factors affecting uptake of Cd in fish. This occurs pri-
tion in liver, muscle, and bone, and appear to be influenced
marily through gill lamellae, and is reduced with increasing
by the amount of fiber in their diet. Lead also accumulates
water hardness (i.e., Ca(II) and Mg(II) concentrations). Two
in eggs and embryos.
mechanisms are thought to be involved in this reduction in
In humans, Pb is initially distributed to various organs
Cd uptake: 1) inhibition of Cd(II) uptake into gill tissue by
and tissues and is gradually redistributed into two compart-
Ca(II), and 2) physiological adaptive responses to increased
ments: an exchangeable compartment, comprising blood
Ca(II) concentrations. After exposure to Cd chloride in wa-
and soft tissues, and a storage compartment, essentially
ter, uptake is greatest in the gill, kidney, and liver. Fish ex-
bone. Lead levels in bone continue to increase throughout
posed to Cd contaminated food accumulate Cd principally
life while stabilizing, or in some cases decreasing, in soft tis-
in the kidney, gut, and liver. One study suggests that reten-
sues (Tsuchiya 1986, see chapter 12 for further details).
tion is greater for Cd present in food than in water (1%
compared with 0.1%). On a subcellular level, Cd has been
found to be distributed both in the cytosol and in the nu-
7.4.5.2. Cadmium
cleus bound to metallothioneins and other Cd-binding pro-
In aquatic systems, Cd is most readily absorbed by organ-
teins rich in cysteine. Biological half-times for Cd in fish
isms directly from the water (primarily over gills and body
have been estimated to range between 24 and 63 days. Birds
surfaces) in its free ionic form (Cd(II)). Although many of its
exposed to dietary levels of Cd chloride accumulate Cd in
inorganic salts are soluble in water (e.g., acetate, chloride,
the liver and the kidney.
and sulfate), they do not appear to be taken up. In fact, Cd
Animal data suggest that absorption over pulmonary sur-
toxicity is generally found to be lower in marine waters, and
faces is higher than gastrointestinal absorption. Depending
this has been attributed to chloride complexation. In marine
on particle size, deposition, and solubility in biological fluids,
waters, Cd is mainly present as soluble chloride complexes.
up to 50% of inhaled Cd may be absorbed, compared with
Cadmium is also easily complexed with some organic com-
(on average) 5% of ingested Cd. The extent of dietary uptake
pounds, which has an important bearing on its bioavailabil-
is reportedly affected by nutritional status, ranging from 1 to
Chapter 7 · Heavy Metals
395
20%. The importance of the protein metallothionein for
the ventral nerve chord corresponded with a general increase
both the distribution and toxicity is described in section
in other tissues (e.g., exoskeleton, foregut, and remaining
12.2.3.2. Additional information can be found in Nordberg
tissues). Mercury levels in the gills remained relatively con-
and Nordberg (1987). Essentially, once in the blood stream
stant. Fiddler crabs (Uca pugilator) are able to transfer Hg
Cd is bound to high and low molecular weight proteins and
from the gills to the hepatopancreas with increasing effective-
distributed around the body. Binding to metallothioneins
ness when temperatures are increased. Few reports are avail-
(low molecular weight components in the blood) invariably
able on methylation or excretion of Hg in arthropods. Blow
leads to transportation to the kidney, where its primary toxic
flies (Lucilia illustris) reportedly tend to eliminate inorganic
effects are experienced. Excretion of Cd from muscle, kidney,
Hg compounds with greater ease than methylated forms.
liver, and the body overall is very slow. In rats, the half-time
Environmental variables such as temperature, pH, and
is 200-700 days, while in humans, half-times reported for
redox potential are particularly important for Hg uptake in
muscle, kidney, and liver range from 10 to 30 years.
fish. Data on uptake suggest that absorption increases with
higher temperatures and lower pH. Tissue concentrations of
Hg increase with age for both marine and freshwater fish.
7.4.5.3. Mercury
Mercury accumulated in fish is usually in the form of methyl-
Background concentrations of Hg are generally low, except
mercury, whereas the source is usually inorganic. It has been
in the immediate vicinity of mining sites, etc. Mercury oc-
proposed that methylation of inorganic forms is facilitated
curs naturally as elemental Hg, and as organic and inorganic
by microbial metabolism, either in the environment prior to
compounds (e.g., Hg vapor, Hg salts, short- and long-chain
assimilation or by microorganisms associated with bronchial
alkylmercury compounds). This speciation of Hg is of great
mucosa. It has also been suggested that inorganic Hg is meth-
importance in relation to uptake from soil and water, as the
ylated in fish liver. Mosquito fish (Gambusai affinis) have been
different chemical forms differ greatly in their physico-chem-
shown to absorb metallic Hg five times faster than inorganic
ical properties. Much of the Hg in the environment is un-
Hg. This is presumably related to the high lipid solubility of
available to organisms, as it is strongly bound to sediment
the metallic form, which allows it to pass through gill mem-
or organic material. Inorganic forms can be methylated by
branes while salts become tightly bound to mucoproteins.
microorganisms and transformed to methylmercury, which
High Hg levels in the gills have been observed in a number of
is much more readily taken up and accumulated in both
species. Methylmercury levels in rainbow trout (Salmo gaird-
aquatic and terrestrial organisms.
neri) steadily increase in the muscles and brain, whereas other
Uptake in aquatic plants increases with increasing con-
Hg compounds accumulate primarily in the kidneys, spleen,
centrations of Hg salts, with greater levels occurring in roots
and liver. In brook trout (Salvelinus fontinalis), this steady in-
than in shoots. However, some of this may be attributed to
crease continued over a long period until a steady-state tissue
differences in surface absorption and not uptake per se. The
level was achieved. During this period, no significant elimina-
possibility of some active uptake has been suggested for ma-
tion was observed. Elimination rates reported for metallic
rine diatoms (e.g., Chaetoceros costatum). A similar pattern
and inorganic Hg (absorbed from water) are similar in mos-
of absorption and distribution is found in studies with ter-
quito fish, having a half-life of 45 days. The half-lives for in-
restrial plants. Soil type strongly influences the extent of up-
organic and organic Hg taken up by the thornback rays (Raja
take. For example, Hg absorption is inversely related to or-
clavata) from diet are 61 and 323 days, respectively.
ganic content. As with aquatic plants, the highest Hg con-
As with fish, birds can assimilate organic forms of Hg
centrations occur in the roots, though there is some evidence
more readily than inorganic compounds. Experiments with
of translocation to other parts of the plant including the
Japanese quail (Coturnix coturnix japonica) indicate that
leaves. It has been suggested that several leguminous plants
uptake is unaffected by route of administration. Methylmer-
appear to be able to eliminate Hg by a process termed `bio-
cury is distributed evenly in tissues while inorganic com-
volatilization' involving the loss of at least 75% of their tis-
pounds accumulated primarily in the liver and kidneys of
sue burden as elemental Hg vapor.
adult birds. The highest concentrations following chronic
Most of the studies of Hg uptake in invertebrates do not
Hg exposure are generally found in the kidneys and liver.
differentiate between external absorption and actual uptake.
Accumulations may also occur in reproductive organs (e.g.,
In the case of methylmercury compounds, uptake usually
goshawks, Accipter gentilis). Excretion appeared to be en-
correlates with surface absorption capacity. Blue mussels
hanced by egg laying, with concentrations of methylmercury
readily accumulate Hg, and have been found to have a bio-
occurring in the egg white and other Hg compounds typi-
logical half-life of 293 days following chronic exposure, and
cally in the yolk. Overall, inorganic forms are more rapidly
only 53 days following acute exposure. Approximately 75%
excreted than methylmercury. Plumage and other keratinized
of the Hg accumulated was immobilized in an inorganic
structures represent an important excretion route for both
form. Another study with a sediment-feeding bivalve (Maco-
forms, though especially for methylmercury.
ma balthica) revealed that only 6% of the total Hg accumu-
The primary uptake route of Hg in marine and terrestrial
lated was present as methylmercury, although this percent-
mammals is though diet. This is related to the relatively high
age was somewhat lower than for mussels in the same area.
concentrations of methylmercury in food items (e.g., fish),
Assimilation of Hg in arthropods depends on both form
which is more effectively taken up than inorganic forms.
(organic or inorganic) and source of exposure, and is highly
Studies with mice, sheep, cows, and humans have shown
variable between taxa (Zauke et al. 1996). When presented
that over 95% of methylmercury given either in drinking
in the diet of shrimp (Hyalella azteca), net absorption of
water or food is taken up. Possible uptake in the lungs and
methylmercury was 70-80%, compared with 38% of inor-
over the skin is possible, but it is unlikely to be of much im-
ganic Hg. Studies of the uptake of radioactive isotopes of
portance due to low environmental levels. Tissue levels have
Hg by grass shrimp (Palaemonetes pugio) indicate that 24
been reported from wild populations of seals, dolphins (Ste-
hours after exposure, methylated Hg is concentrated mainly
nella coeruleoalba), and common or harbour porpoise (Pho-
in the ventral nerve chord, whereas assimilated inorganic Hg
coena phocoena) (Andre et al. 1990). Highest levels gener-
is found predominantly in the gills. However, after 72 hours
ally occur in the liver. Levels present in bone tissue are be-
the tissue distribution changed; concentration decreases in
lieved to be correlated with age.
396
AMAP Assessment Report
such as dimethylselenide, can also be excreted from the body
7.4.5.4. Selenium
by exhalation. Experiments with rats, however, have demon-
The metalloid Se is widely distributed in the environment,
strated that this occurs only with dosages approaching lethal
existing naturally in the form of selenide (Se (II)), elemental
concentrations.
Se, selenite (Se (IV+)), and selenate (Se (VI+)). Most Se com-
Two elimination phases are evident from animal studies.
pounds are water soluble and are thus readily available for
In the first rapid phase, the biological half-life of Se ranges
dermal and gastrointestinal uptake. Under acidic conditions
from one to three days, depending on the compound and
or in biological systems, Se compounds may be reduced to
dose. The second phase of elimination, which has been
selenide, which may replace sulfur in amino acids (e.g., se-
found to be independent of dose, ranges from 30 to 70 days
lenomethionine and selenocysteine). Many Se compounds
in most species. This pattern is slightly different in humans,
are volatile and relatively unstable at room temperature
where three phases have been identified with biological half-
(e.g., hydrogen selenide). Under certain conditions, there-
lives of about 1 day, 8-20 days, and 65-116 days (Högberg
fore, the lungs can play an important role in the uptake of
and Alexander 1986).
soluble as well as insoluble Se compounds.
Little or no information is available on the characteristics
of uptake, distribution, and excretion of Se in plants, inverte-
7.5. Toxicological effects
brates, fish, and birds. Limited data are available for mam-
mals from laboratory studies with rats, mice, dogs, mon-
Numerous laboratory experiments have been carried out to
keys, and ruminants. Certain plant groups accumulate Se to
assess the toxicological effects of acute and chronic doses of
very high levels when grown in highly contaminated soils
heavy metals. Intoxicants are often administered via inhala-
(e.g., several species of Astragalus, and some Haplopappus
tion, ingestion, injection, or absorption through the skin, so
and Stanleya), whereas other plants (e.g., Aster, Atriplex,
that effects can be directly related to dosage, forming the
Castelleja, Grindelia, Gutierrezia, Machaeranthera, and
basis of dose-effect and dose-response relationships (Pfitzer
Menzela) concentrate Se to lesser levels. Only low-level ac-
and Vouk 1986). Most experiments focus on clinical signs
cumulation occurs in grasses, grains, and weeds. The ability
and symptoms of lethal and sublethal toxicity. In contrast to
of plants to accumulate Se from soils with low Se content is
this, the data currently available for heavy metal levels in tis-
also variable (Högberg and Alexander 1986).
sues of Arctic organisms are generally obtained from envi-
Results from animal studies suggest that Se compounds
ronmental monitoring of natural populations. In this case,
are effectively absorbed in the gastrointestinal tract. The
the emphasis is on exposure rather than dosage, and the
rates of absorption were 90% for selenite and 80% for se-
focus is on the circumstances and exposure levels that may
lenomethionine or selenocysteine administered orally. The
elicit effects (Elinder 1984). In the Arctic, metal concentra-
degree of uptake was found to be much lower in sheep and
tions, and hence dosages, generally do not change quickly
cows (30-35%), although this probably reflects the reduc-
(except in the event of a spill or an accidental release), and
tion of selenite to its elemental form by bacterial action in
therefore information in relation to acute toxicity is less ap-
the gastrointestinal tract (Högberg and Alexander 1986). Al-
plicable than information relating to chronic toxicity.
though little is known of the physiological processes govern-
Although many reliable measurements of metal levels in
ing absorption of simple Se compounds, there is some
organisms have been obtained, there are relatively few mea-
evidence that selenomethionine can be transported against
surements of metals in the environment in which the organ-
a concentration gradient. Corresponding data for Se uptake
isms live. Consequently, in order to evaluate the effects of
in the lungs are unavailable. In the lungs of rats, uptake of
heavy metal exposure on Arctic biota, it is necessary to ex-
selenious acid (H2SeO4) was 94%, and elemental Se, 57%.
trapolate from results obtained under specific, laboratory
No data are available for dermal uptake, although it has
exposure conditions. As little experimental toxicology has
been observed in rats.
been carried out with Arctic organisms to date, in the major-
Once absorbed, water soluble Se is rapidly distributed to
ity of cases conclusions have to be drawn from counterparts
most major organs, with highest concentrations in the liver
in temperate climates. The extent to which results obtained
and kidney. Little is known about Se transport. It has been
for temperate species are applicable to Arctic species is not
suggested that after accumulation in red blood cells, Se com-
entirely clear, especially in light of the intrinsic differences
pounds are translocated to plasma proteins (selenoproteins),
between species in terms of sensitivity and responses to
which in turn may play an important role in transporting Se
heavy metal exposure. Furthermore, environmental condi-
to the liver. In humans, Se binding to plasma lipoproteins
tions associated with changing seasons are more extreme in
has been demonstrated. The form of Se has been observed
polar regions, and there has been almost no experimental
to have a marked influence upon accumulation; selenite-de-
study of how seasonal variations influence toxicity.
rived Se accumulates more rapidly in liver and kidney than
The effects of heavy metal intoxication are many and var-
selenate-derived Se (Högberg and Alexander 1986).
ied, depending on the metal, the organism involved, the route
Mammals appear capable of maintaining Se homeostasis
of uptake, metabolism or biotransformation, and rates of
at low levels of exposure. Many Se compounds are biotrans-
elimination. The following is a brief overview of informa-
formed within the liver to excretable metabolites (e.g., meth-
tion on this subject contained in the WHO's International
ylated selenides such as trimethylselenide and dimethylse-
Programme on Chemical Safety (IPCS). This information in-
lenide). Studies on rats have revealed that the predominant
cludes the typical characteristics of and toxic effects attribut-
excretory pathway is via urine; the fraction excreted de-
able to each of the four heavy metals of greatest concern in
pends on nutritional status and dosage. At high levels, up to
the Arctic environment, Pb, Hg, Cd, and Se. Where possible,
67% of Se dosage is excreted in urine compared with 10%
results from studies relating heavy metal burdens to toxicity
in feces (via bile). In contrast to the general pattern, rumi-
will be cited for possible comparison with levels reported in
nants can excrete up to 66% of the total Se dosage in feces.
Arctic flora and fauna (see section 7.7.3 and Table 7·22).
This increased fecal excretion has been attributed to poor
Although there are numerous data available for heavy
gastrointestinal absorption rather than a greater rate of
metal concentrations in various Arctic biota tissue, there are
elimination. Significant quantities of methylated metabolites,
few reports of toxic effects of heavy metals in relation to
Chapter 7 · Heavy Metals
397
actual tissue levels. Most of the data contained in the IPCS
acetate, egg production in Japanese quail was depressed, and
regarding heavy metal toxicity are presented in the form of
this effect was enhanced with increased levels. Where dos-
experimental concentrations (i.e., lethal concentrations
ages were high, egg production was almost completely de-
LC50) or dosage administered (i.e., lethal dose LD50),
pressed and the eggs which were produced were either soft
rather than resulting tissue burdens. Therefore, in the ab-
shelled or without shells. Hatch rates were also reduced.
sence of other data, only potential effects of heavy metal
Birds can be exposed to high dosages of metallic Pb
intoxication can be described below.
through the ingestion of pellets of lead shot. Over a period
of 20 days, adult mallard ducks showed clear symptoms of
Pb poisoning in response to sublethal dosages, including
7.5.1. Lead
green diarrhea, anorexia, and weakness. Short-term Pb toxi-
According to WHO (1989a), Pb levels present in the envi-
cosis involving varying degrees of paralysis and abnormal
ronment are unlikely to affect aquatic plants. Reductions in
locomotor function has been reported as being transitory,
photosynthesis and respiration have been reported in sun-
disappearing after eight days. Intoxication with metallic Pb
flowers along with reductions in overall growth in lettuce
also brought about chronic inhibition of erythrocyte delta-
and carrots after exposure to very high Pb concentrations.
ALAD activity. Results also indicate that trialkyllead com-
In was concluded that Pb is only likely to affect terrestrial
pounds are very toxic, but effects have only been reported
plants at very high environmental concentrations (e.g.,
for starlings (Sturnus vulgaris). The symptoms included se-
> 0.005 mg/L).
vere lack of co-ordination, resulting in decreased feeding.
The tolerance of invertebrates to Pb salts is variable, par-
In mammals, the major effects of Pb are related to the
ticularly where populations have experienced prior expo-
hematopoietic, nervous, gastrointestinal, and renal systems.
sure. Symptoms of toxicity reported for nematodes and ca-
Symptoms and effects of inorganic Pb poisoning are de-
terpillars include impaired reproduction and development.
scribed for humans in chapter 12. It is assumed that the
Caterpillars (Scotia segetum) fed contaminated food through-
principal effects of Pb intoxication in mammals are likely to
out development showed decreased survival to adult stage,
be similar to those described for humans. The primary ef-
and those that did survive were visibly deformed and unable
fects of Pb intoxication include 1) interference with the syn-
to produce eggs. Mature mussels were able to tolerate large
thesis of red blood cells which are necessary for oxygen
amounts of Pb in their tissues without apparent toxic effects
transport (anemia), 2) damage to both the central and per-
(by enclosing the metal in membrane-bound vesicles). Settle-
ipheral nervous system leading to encephalopathy and neu-
ment of oyster larvae was found to be significantly reduced
rological dysfunction, and 3) gastrointestinal symptoms such
for Crassostrea gigas, as exposure to > 0.01 mg/L Pb brought
as loss of appetite, diarrhea, constipation, and in severe cases
about reduced larval development and therefore delayed
colic (Tsuchiya 1986).
peak settlement time. Reduced oxygen consumption has
been reported for shrimp exposed to high concentrations of
7.5.2. Cadmium
Pb. A dose-related decrease in oxygen consumption (0-2.0
mg/L) was observed with crayfish (Orconectes virilis) ex-
Few reports are available concerning the toxic effects of Cd
posed to Pb acetate for a period of ten days, although long-
on plants. Reduced growth and the loss of chlorophyll-a
term acclimatization was evident after 20, 30, and 40 days.
from leaves have been reported for two floating weeds
Although Pb apparently reduces the capacity for oxygen up-
(Lemna minor and Salvinia natans) and water hyacinth
take through the gills, the crayfish were able to compensate
plants exposed to Cd. Plants grown in soil are generally in-
by increasing the flow of water over gill surfaces. Growth
sensitive to Cd except at high concentrations. Two tree spe-
suppression has been reported for larval brine shrimp (Arte-
cies, white pine (Pinus strobus) and red maple (Acer rubrum),
mia) with exposure to concentrations exceeding 5 mg/L.
responded differently to high Cd exposure. Red maple seed-
Chronic exposure of adult fish to inorganic Pb can cause
lings developed interveinal chlorosis and stunting of the
sublethal effects on morphology, enzyme activity, and avoid-
leaves, wilting, and death, while pines experienced inhibition
ance behavior. Juvenile stages are more susceptible to Pb
of needle expansion. Effects were also observed on root de-
than adults or eggs. In a study designed to establish accept-
velopment. At high dosages, there was both a reduction in
able toxicant limits for inorganic Pb, Davies et al. (1976,
the number of new roots and a stunting of those already
cited in WHO 1989a) observed blackening of the tail, fol-
present. Stunted growth and similar toxic symptoms have
lowed by spinal curvature and eroded caudal fins in re-
also been observed in lettuce, cabbage, carrot, and radish
sponse to exposure to sublethal concentrations of Pb. These
plants. Physiological studies have demonstrated that increas-
sublethal concentrations were not related to tissue burdens.
ing Cd concentrations reduce stomatal opening, leading to a
Other studies have also reported anemia and basophilic
reduction in both transpiration and photosynthesis.
stippling of erythrocytes. The presence of Pb is also believed
Laboratory studies of acute toxicity suggest that Cd is
to result in reduced larval survival due to skeletal malforma-
moderately to highly toxic to aquatic invertebrates, depend-
tion, erosion of fins, outgrowths from the fry, and poor ab-
ing on a variety of environmental variables (temperature,
sorption of yolk.
salinity, organic content of water, and chelating agents). Ef-
A number of studies reporting Pb toxicity in birds, main-
fects of long-term exposure can include larval mortality and
ly gallinaceous species, have revealed that both Pb salts and
temporary reduction in growth. In an investigation of sub-
metallic Pb are not toxic except at very high concentrations.
lethal concentrations and effects on adult shrimp (Palae-
Comparable sensitivities are assumed for other bird species.
monetes pugio), Cd body burdens of 40 g/g were found to
For chickens, Japanese quail, and mallard ducks (Anas platy-
inhibit molting. At more moderate concentrations of 23
rhynchos) a variety of sublethal acute effects have been re-
g/g and 10 g/g, molting was stimulated. Histological ef-
ported, including lethargy, weakness followed by anorexia
fects at near-lethal concentrations included blackening and
and anemia, and an overall reduction in weight gain. High
damage to gill filaments, although no body residue levels
Pb levels in blood have been correlated with suppression of
were given. Terrestrial invertebrates, on the other hand, are
the activity of the enzyme delta-ALAD in erythrocytes,
relatively insensitive to Cd exposure, probably due to the
which is involved in heme synthesis. With low doses of Pb
presence of effective sequestration mechanisms in specific
398
AMAP Assessment Report
organs. Nevertheless, nematodes exposed to sublethal doses
from the intestine. Cadmium is seldom the cause of liver
of Cd do not grow as large as controls and have significantly
damage in mammals; however, experiments with rats involv-
reduced reproductive success. Similarly, growth and repro-
ing long-term exposure (six months), with final Cd concen-
duction are the parameters found to be most sensitive to Cd
trations of 188 g/g, caused changes in liver cell morphol-
in the ollembolan (Orchesella cincta) and the oribatid mite
ogy and enzyme activity (Elinder 1985).
(Platynorthrus peltifer). Few studies are available on the
toxic effects of Cd on mollusks. Some work has been carried
7.5.3. Mercury
out on effects on early development stages of the mud snail
(Ilyanassa obsoleta) and the sea star (Asterias rubens). In
Susceptibility to toxic effects from high Hg accumulation
both cases, development was adversely affected. Subadult
varies greatly. The degree of toxicity is related to a number
garden snails fed Cd responded with decreasing food con-
of environmental parameters including temperature, salinity,
sumption with increasing dosage. At high levels, both growth
dissolved oxygen, and water hardness. Organic forms of Hg
and reproductive activity were suppressed.
are generally more toxic than inorganic forms in both aquatic
Acute toxicity of Cd has been studied in a variety of fish
and terrestrial organisms. This is related to the lipophilic
with respect to lethal concentrations. LC50 values have not
properties of fine Hg compounds. As a protective mechanism,
been related to tissue burdens. Sublethal effects have been
some microorganisms biotransform inorganic Hg complexes
reported for a number of fresh water and saltwater species.
assimilated from their surroundings into organic forms. This
Eggs and larvae are most sensitive to Cd exposure, respond-
process of biomethylation renders Hg much more accessible
ing to exposure with either delayed hatching or non-emer-
to higher organisms.
gence and inferior larval growth. The most sensitive indica-
A number of adverse effects have been reported for plants
tor of Cd toxicity in Atlantic salmon (Salmo salar) was
exposed to a wide range of Hg concentrations. High concen-
found to be inhibition of growth of alevins. When three
trations of inorganic Hg affect macroalgae (Laminara hyper-
generations of brook trout were intoxicated, significant
borea) by reducing gametophytic germination. Adverse ef-
numbers of first and second generation males died during
fects on the growth of the red algae (Plumaria elegans) have
spawning. Cadmium toxicity in fish is characterized by ionic
been reported, with reductions of up to 50% following acute
imbalance with reduced plasma Ca(II), Na(II), and Cl. The
exposure. Significant reductions in growth have also been ob-
probable mechanism underlying this toxicity is inhibition of
served in cauliflower, lettuce, carrots, and potato tubers. Wa-
ion-transporting enzymes by Cd(II) present in gill mem-
ter cabbage (Pistia stradiotes) exposed to various concentra-
branes. Furthermore, Cd has been shown to inhibit Na/K-
tions of inorganic Hg responded by increasing levels of free
ATPase in the gills, thereby having an influence on ATP pro-
amino acids and decreasing chlorophyll content, protein,
duction in the gill. Chronic exposure may also bring about
RNA, and enzymatic activities (catalase and protease). These
prolonged hypoglycemia and reduced ability to absorb glu-
changes typically lead to senescence. A closer look at the tox-
cose and fructose in the gut. Cadmium has also been shown
ic effects at the tissue level in Elodea densa revealed differen-
to interfere with calcium metabolism. Minnows (Phoxinus
ces between tissues. Apical cells and roots developed abnormal
phoxinus) exposed to various concentrations of Cd devel-
nuclear and mitochondrial characteristics, although roots
oped fractured vertebrae in the caudal region. Spinal abnor-
were less sensitive. Mitotic activity was inhibited in root meri-
malities were also observed with newly hatched medaka fry
stems. In the presence of methylmercury, mitosis was pro-
(Oryzias latipes) after eggs were exposed. Fish showing mal-
moted in bud meristems, although cell divisions were abnor-
formations of the spine after exposure to Cd had significant-
mal. Methylmercury inhibited the development of both root
ly less Ca(II) in the vertebral column than did control fish.
and bud initials in Elodea densa. Finally, a study with roots
Kidney damage has been reported in wild colonies of
and shoots of barley plants indicates that inorganic Hg may
pelagic sea birds having Cd residues of 60-480 g/g. Similar
also significantly reduce K+ and PO4 uptake and translocation.
results have also been obtained in laboratory studies in
There are no reports of the effects of chronic exposure on
which birds were dosed so that Cd levels in kidney ranged
invertebrates. Most studies have focused on acute and short-
from 95 to 240 g/g. The damage included cell necrosis, nu-
term effects. Invertebrates in general are more susceptible to
clear pyknosis, mitochondrial swelling, and some tubulor-
Hg toxicity during the larval stages. Acute exposure inhibits
rhexis, although there was some indication of regeneration.
embryogenesis of developing American oyster larvae. Studies
Spermatogenesis was reduced in wood pigeons (Columba
of the effects of Hg on crustacean larvae have revealed ex-
palumbus) and mallards exposed to Cd.
tended periods of metamorphosis and decreased survival
Chronic Cd exposure produces a wide variety of similar
(blue crabs, Callinectes sapidus), as well as reduced settling
acute and chronic effects in mammals and humans. These
(barnacles, Balanus balanoides). Larval shrimp (Palaemo-
are discussed in detail in section 12.2.3.2. Kidney damage
netes vulgaris) did not show immediate responses to acute
and lung emphysema are the primary effects of high levels
exposure. Instead, delayed effects were noted including re-
of Cd in the body. No laboratory investigations have been
duced post-larval survival, delayed molting, extended devel-
carried out on the toxic effects of high Cd burdens in Arctic
opment times, increased number of larval instars, and an
mammals. However, based on human toxicology, it is rea-
increase in the occurrence of abnormalities.
sonable to assume that Cd concentrations of 100-200 g/g
There is much less information available on potential ef-
in the kidneys may represent a risk. At these concentrations,
fects in adult invertebrates. Significant reductions in growth
the first signs of renal damage (proteinuria) have been ob-
have been observed in some species (e.g., blue mussels and
served in land mammals and humans. Accumulation of Cd
terrestrial slugs). Adult crabs (Carcinus maenus) exposed to
in the kidney can also cause disturbance in Vitamin D and
high Hg concentrations displayed reduced cardiac activity
calcium metabolism, and in severe cases can lead to osteo-
and oxygen consumption, and in severe cases a loss of the
malaci and osteoporosis. Experiments with rats have shown
ability to osmoregulate. Mercury is also believed to decrease
that Cd concentrations in the range of 50-170 g/g can
the rate at which limb regeneration occurs in crabs (e.g.,
cause an inhibition in the change of Vitamin D to the active
Uca pugilator). A single study of behavioral responses of
form 1,2,5 DHCC, which has an important regulating effect
grass shrimps (Palaemonetes pugio) to acute intoxication
on calcium metabolism and stimulates the uptake of calcium
demonstrated that the probability that prey organisms would
Chapter 7 · Heavy Metals
399
be caught by predators increased after exposure, indicating
logical level, effects have been described as resulting from
a general decrease in responsiveness.
degeneration of cerebellum and medulla oblongata and from
Physiological biochemical abnormalities have been de-
demyelination of the spinal cord and peripheral nerves. In
scribed for a number of fish species following exposure to
red-tailed hawks, observed effects on feeding and eventual
sublethal concentrations of Hg. These include depressed ol-
weight loss were also partly attributable to feeding difficul-
factory response with increasing concentrations and expo-
ties resulting from impaired muscle co-ordination. However,
sure duration (e.g., rainbow trout and Anabas scandens);
autopsies of goshawks exposed to Hg revealed that the dom-
reduced gill filament respiration due to damage to secondary
inant effect leading to weight loss was likely to be muscular
lamellae (e.g., roach, Leuciscus rutilus); blindness; reduc-
atrophy. In many cases, reductions in egg production and
tions in hemoglobin content, erythrocyte count, body weight,
hatchability have also been reported. Some studies have de-
and body protein (Anabas scandens); and, finally significant
scribed significantly reduced egg production, chick weight,
decreases in intestinal absorption rates of glucose, fructose,
and survival. In addition, large numbers of thin-shelled and
glycene, and trytophan. Osmoregulatory effects have also
shell-less eggs were seen in some studies (e.g., penned pheas-
been observed in some fish (rainbow trout), particularly in
ant, Japanese quail). Mercury concentrations above 30 µg/g
response to exposure to inorganic Hg. It has been suggested
ww in liver and kidney and 2-3 µg/g ww in eggs may result
that this symptom arises as a result of enhanced mucus pro-
in detrimental effects on birds (see Table 7·22).
duction on the gills caused by Hg exposure. There is also
Although Hg is primarily a neurotoxin in mammals, it is
some suggestion that immune response can be affected.
also responsible for damaging reproductive capacity as it in-
In addition to the effects already described for fish, Hg
terferes with spermatogenesis, causing chromosome damage
has an adverse effect on reproductive performance in many
in connection with mitosis. Developing fetuses, and breast-
species. Chronic exposure was seen to suppress oocyte de-
feeding offspring in particular, are sensitive to Hg in the form
velopment in freshwater teleosts (e.g., Channa punctatus).
of methylmercury. WHO has concluded that the developing
This effect corresponded with the suppression of gonado-
fetus is especially susceptible to Hg due to intense cell divi-
trophs in the pituitary, inducing a gonadal `resting phase'.
sion and formation of the central nervous system. An effec-
Furthermore, studies investigating the effects of exposure
tive placental barrier exists against the transfer of inorganic
on rainbow trout sperm indicate that Hg in the water col-
Hg, but not methylmercury which readily transfers from
umn has the potential to significantly decrease fertilization.
maternal to fetal tissue and accumulates to some extent in
Environmental contamination also has the potential to in-
the fetal brain (Bremner 1974).
hibit hatching success (e.g., zebra fish, Brachydanio rerio)
As outlined in Table 7·22 (section 7.7), liver and kidney Hg
and promote embryo deformation. Killifish (Fundulus hete-
concentrations above 25-60 µg/g ww may constitute a risk for
roclitus) exposed to inorganic Hg during embryonic devel-
Arctic mammals. Experiments on other animals have shown
opment showed an increasing incidence of spinal curvature,
that the organs most susceptible to damage from methylmer-
although the susceptibility of eggs was variable.
cury and inorganic Hg are the central nervous system and kid-
Most of the effects identified for inorganic Hg intoxica-
ney, respectively. There are no guideline concentration limits
tion in birds have been reported for gallinaceous species,
available for these organs in humans, as the parameter used
which are considered to be unrepresentative of other birds.
for determining Hg burden is blood concentration. Marine
In general, organic Hg is more toxic than inorganic salts.
mammals with Hg levels in the blubber could potentially be
The primary effect, which is evident from various studies
at risk at times when food is scarce. However, concentrations
(e.g., chickens and Japanese quail) is that birds refuse both
in blubber of Arctic marine mammals are generally low.
food and water, leading to subsequent poor growth. There-
fore the clinical effects may be both direct and indirect. It
7.5.4. Selenium
has been estimated that a period of at least two weeks of ex-
posure is required before the direct effects are likely to occur.
Overall, there is little information on the toxicity of Se.
A number of more subtle effects have also been identified
There is no information on the effects of acute and chronic
from laboratory studies, including effects on enzyme systems,
exposure to plants, invertebrates, fish, or birds. Limited data
cardiovascular function, blood parameters, immune response,
are available for animals and humans. Under acute expo-
and kidney function and structure. The activity of two en-
sure, the critical organ of Se toxicity appears to be the cen-
zymes in particular, cholinesterase and lactate dehydrogenase,
tral nervous system. Pathological studies on animals suffer-
can be altered by oral Hg exposure in Japanese quail. Cholin-
ing from acute Se poisoning have revealed liver and kidney
esterase activity decreased with dosage while lactate dehydro-
congestion, endocarditis, myocarditis, and peticheal hemor-
genase activity increased to three times its original level irre-
rhages of the epicardium. Selenium intoxication in dogs and
spective of dose. In chickens, Hg intake induced cardiovascu-
rats reportedly induces vomiting, dyspnea, tetanic spasms,
lar disturbances in the form of myocardial changes and alter-
and finally death by respiratory failure. In a study investigat-
ations in the number of red blood cells, hematocrit, and cor-
ing the long-term effects of Se, rats fed on seleniferous se-
puscular volume. Chronically exposed chickens showed sup-
same meal suffered from hepatic lesions with 7.34 g/g Se in
pressed primary and secondary immune responses. Damage
the liver, whereas rats with 0.72 g/g had no lesions.
to the proximal tubules of the kidney has also been reported
Toxic effects resulting from long-term Se exposure through
to result from oral exposure (e.g., juvenile starlings). Some re-
consumption of accumulatory plants have been reported for
productive effects have been reported (e.g., reduced fertiliza-
livestock. The so-called `Blind staggers' syndrome, which
tion and weakening of the egg shell), although these are not
can take several weeks to develop, is characterized by im-
as prominent for inorganic Hg salts as for organic Hg.
paired vision, reduced appetite, and a tendency to wander.
Sublethal doses of organic Hg result in neurological im-
The effects of chronic poisoning on livestock (the so-called
pairment, reproductive effects, and weight loss in birds. Neu-
`alkali disease') include emaciation, hair loss from the mane
rological symptoms including impaired co-ordination of mus-
and tail, deformation and shedding of hooves, and erosions
cle movements and ataxia have been observed in a number of
of the joints of long bones. Liver cirrhosis may develop in
birds (domestic fowl; penned pheasant, Phasianus colchicus;
severe cases. Similar effects have also been reported in sheep
red-tailed hawk, Buteo jamaicensis; goshawk). On a physio-
and dogs (Högberg and Alexander 1986).
400
AMAP Assessment Report
Anecdotal reports suggest that excess Se may also de-
tain these laminar structures, although there is often consid-
crease reproductive performance in farm livestock and
erable wind shear between the layers (Radke et al. 1989).
hatchability in chickens. This has also been reported to be
Aerosol concentrations in the lower troposphere show sig-
the case with mice. In fact, Se accumulates in mammalian
nificant variations depending on meteorological conditions,
testes to a greater extent than in other tissues (see review by
primarily temperature inversions.
Hansen and Deguchi 1996), and influences spermatozoal
The measurements of vertical profiles revealed layers of
motility. Selenium is incorporated in the mitochondrial cap-
particles in the upper part of the High Arctic troposphere,
sule of sperm, and in this way is likely to affect its behavior
although particle concentrations were much lower than those
and function. In humans, both high and low levels have
in the lower troposphere. Aerosol size measurements in the
negative effects on the number and motility of spermatozoa.
lower (up to 2-3 km) and upper (from 3 to 4-5 km) parts of
No data are available for the effects of Se on female fertility.
the troposphere during winter were highly variable, with
Studies of this type are difficult to interpret and extrapo-
fine particles (< 1.0 µm) dominating the lower part and
late to other mammal species. Two independent studies
coarser particles in the upper part. The larger concentrations
have demonstrated that, under certain circumstances, Se
of small particles in the lower part of the troposphere ap-
compounds are less toxic to animals which normally have
peared to be associated with air masses transported directly
a high dietary intake of Se.
from emission regions. They may also have been a result of
gas-to-particle conversion (Bodhaine 1989).
During summer, aerosol concentrations are usually very
low throughout the High Arctic. However, vertical profiles
7.6. Regional and circumpolar levels
measured in the Norwegian Arctic sporadically indicated
layers of polluted air with a well defined lower boundary at
and trends of metal contaminants
about 2 km (Pacyna and Ottar 1988). Particles in the upper
7.6.1. Atmosphere
part of the troposphere in summer were of similar size to
7.6.1.1. Air concentrations in the High Arctic
those measured in winter.
Air contamination from anthropogenic sources has been
Fine particles contain high concentrations of several an-
measured throughout the Arctic air mass (Heidam 1984,
thropogenic heavy metals, particularly during episodes of
Lowenthal and Rahn 1985, Ottar et al. 1986, Shaw 1988,
long-range transport of air masses passing over the major
Barrie et al. 1989, Djupstrøm et al. 1993) (Figure 7·18). Ver-
emission source regions (Ottar et al. 1986). Coarse particles,
tical profiles of air concentrations obtained during several
not correlated with haze, consist mainly of clay minerals,
research programs have added to the understanding of the
other soil constituents, and, to a lesser extent, seasalt com-
physical and chemical characteristics and origin of Arctic air
pounds. The chemical composition of particles measured in
contamination. Of particular interest in this connection are
the Norwegian Arctic during a winter flight is shown in Fig-
the results of the three Arctic Gas and Aerosol Sampling
ure 7·22. Data obtained on this and other research flights in
Program (AGASP) campaigns in 1983 (Schnell 1984), in
the Norwegian Arctic (Pacyna et al. 1985) indicate the pres-
1986 (Herbert et al. 1989) and in 1989 (Schnell et
ence of high concentrations of several anthropogenic heavy
al. 1991); the 1988 Canadian Polar Sunrise Experiment
metals and natural compounds in the lower layer of the Arc-
(Barrie 1991); and the British Petroleum (BP) program in the
tic troposphere (up to 2-3 km). The upper layer (from 3 to 5
Norwegian Arctic from 1982 to 1984 (Ottar et al. 1986).
km) contains a mixture of natural compounds and anthro-
The following description of Arctic air pollution is based on
pogenic heavy metals with the latter group in small concen-
measurements carried out within these programs. In the
trations.
lower troposphere (up to 3 km) particle layers are frequently
Several ground-measurement campaigns have been car-
strongly banded (Ottar and Pacyna 1986, Shaw 1986). The
ried out in the High Arctic to assess the level of heavy metals
temperature and wind profiles measured through these lay-
and other chemicals and to use these data to identify the
ers suggest that their thermal stability is sufficient to main-
origin of Arctic air pollution. Knowledge of the subject ex-
Altitude
m
Cl
Sc
Ti
V
Mn
Fe
Co
Cu
Zn
As
Br
Cd
Sb
Sm
Au
4 800
4 200
3 600
3 000
2 400
1 800
1 200
600
0
3.5 7
4 8
7 14
7 14
3.5 7
1.7 3.4
3.5 7
7 14
3.5 7
3.5 7
3.5 7
3.5 7
3.5 7
7 14
1 2
-2
-2
-1
2
-1
2
-1
-1
-3
-1
x 10
x 10
x 10
x 10
x 10
x 10
x 10
x 10
x 10
x 10
x 10
x 10
ng/m3
Figure 7·22. Composition of particles measured at different altitudes during winter flights in the Norwegian Arctic. (After Pacyna et al. 1985).
Chapter 7 · Heavy Metals
401
ng/m3
Table 7·7. Winter and summer median concentrations (ng/m3) of heavy
metals at Ny-Ålesund (Svalbard) and Alert (Canadian Arctic) in 1984.
10
Ny-Ålesund
Ny-Ålesund a
Alert b
8
Winter:
Winter:
Summer
Summer
Metal
Winter
Summer
ratio
Winter Summer
ratio
6
Cu
0.90
0.30
3.0
1.52
0.55
2.8
Ni
0.38
0.10
3.8
0.25
0.07
3.6
Pb
1.76
0.20
8.8
2.15
0.22
9.8
4
Zn
2.30
0.15
15 .0
3.88
0.97
4.0
a. Data from Norwegian Institute of Air Research (NILU).
2
b. Data from L. Barrie pers. comm.
vious winter campaigns, but for some heavy metals having
0
strong anthropogenic sources (such as V, Mn, Ni, Zn, As,
V
Ni
Zn
As
Pb
Mo, Sb, and Pb), they are clearly lower.
Much less information exists on heavy metal concentra-
1983
1984
1986
1987
1989
tions during summer. A comparison of median concentra-
Figure 7·23. Atmospheric winter concentrations (median values) of metals
tions measured at Ny-Ålesund and Alert in the Canadian
measured in different years in the < 2.5 µm aerosol fraction at Ny-Ålesund,
Arctic in winter and summer of 1984 are shown in Table 7·7
Svalbard. (After Maenhaut et al. 1996).
(Maenhaut et al. 1989). The ratios of winter concentrations
panded greatly after North American and Western Euro-
to summer concentrations range from 3 to 15. The ratios for
pean research agencies started a coordinated research net-
individual metals are generally similar for the Alert and Ny-
work in 1977. This effort resulted in extensive data sets of
Ålesund locations.
concentrations in surface air for Greenland (Heidam 1984),
Results of the three concurrent measurement campaigns
the Norwegian Arctic (Larssen and Hanssen 1980, Heint-
at Poker Flat (PKF) and Point Barrow (BRW) in Alaska and
zenberg et al. 1981), northern Canada (Barrie and Hoff
Ny-Ålesund carried out in 1986 have been compared (Djup-
1985, Barrie et al. 1989), and northern Alaska (Rahn and
strøm et al. 1993). Data for As and V are presented in Fig-
Heidam 1981, Rahn et al. 1981, Shaw 1988). Since then,
ure 7·24. It was concluded that polluted aerosols trans-
information on heavy metals in the background aerosol in
ported to the High Arctic within air masses from source re-
the Russian Arctic has become available (Vinogradova et al.
gions at lower latitudes can be trapped in the polar region,
1993).
particularly during winter. Having rather low potential for
Median winter concentrations of heavy metals and other
deposition due to their small particle size, aerosols can circu-
compounds measured in 1983, 1984, 1986, 1987, and 1989
late with the Arctic air masses over the region. Therefore, a
at Ny-Ålesund, Spitsbergen, are presented in Figure 7·23
load of air pollution originating, for example, in northern
(after Maenhaut et al. 1996). The median values in the 1989
Eurasia can under certain conditions reach Alaska first and
campaign are generally of the same order as those in the pre-
then travel farther in a return flow over the North Pole to
Point Barrow, Alaska
Poker Flat, Alaska
Ny-Ålesund, Svalbard
Poker Flat
Point Barrow
Ny-Ålesund
As
2.6
As
ng/m3
ng/m3
2
2
1
1
0
0
2 - 04 - 86
3 - 04 - 86
4 - 04 - 86
5 - 04 - 86
6 - 04 - 86
7 - 04 - 86
8 - 04 - 86
9 - 04 - 86
10 - 04 - 86
11 - 04 - 86
12 - 04 - 86
13 - 04 - 86
14 - 04 - 86
15 - 04 - 86
16 - 04 - 86
V
5.9
V
ng/m3
ng/m3
4
4
3
3
2
2
1
1
0
0
2 - 04 - 86
3 - 04 - 86
4 - 04 - 86
5 - 04 - 86
6 - 04 - 86
7 - 04 - 86
8 - 04 - 86
9 - 04 - 86
10 - 04 - 86
11 - 04 - 86
12 - 04 - 86
13 - 04 - 86
14 - 04 - 86
15 - 04 - 86
16 - 04 - 86
Figure 7·24. Concentrations of As and V in aerosols measured during concurrent measurement campaigns on Svalbard and two locations in Alaska dur-
ing April 1986. (After Djupstrøm et al. 1993).
402
AMAP Assessment Report
1980
1990
ng/m3
ng/m3
5
2
4
ng/m3
ng/m3
1
1
3
6
0
0
2
5
Point Barrow
Poker Flat
1
ng/m3
6
4
0
ng/m3
3
6
5
Wrangel Island
5
4
2
4
1
3
3
0
ng/m3
2
2
Severnaya Zemlya
4
1
1
3
0
0
2
Alert
Alert
1
ng/m3
ng/m3
2
0
ng/m3
2
3
1
Ny-Ålesund
1
2
0
0
Nord
1
Nord
0
Ny-Ålesund
1980
1990
European emissions,
European emissions,
tonnes
tonnes
100 000
100 000
80 000
80 000
60 000
60 000
40 000
40 000
nc nc
As Pb Zn Cu
Ni
20 000
Mn V
20 000
0
0
Figure 7·25. Winter concentrations of metals in air at remote locations in the Arctic at the beginning of the 1980s and 1990s, and European emissions
of As, Pb and Zn during the same periods. (Data sources: Ny-Ålsund, Norwegian Institute of Air Research; Alert, Len Barrie pers. comm.; Nord, Niels
Heidam pers. comm. and Danish Environmental Protection Agency 1997; Severnaya Zemlya and Wrangel Island, Russian State Committee for Hydro-
meteorology Assessment `Heavy metals in the Arctic'; Point Barrow and Poker Flat, Glen Shaw pers. comm. and (Pb, Zn, Ni from `Gates of the Arctic'
station, Fairbanks near Poker Flat) Polissar et al. 1996).
the Norwegian Arctic. In this way concentrations measured
7.6.1.2. Concentrations of heavy metals in subarctic air
at various stations in the High Arctic are intercorrelated. In-
deed, analysis of a 4-year chemical data set from Alaska
Air concentrations of heavy metals in the High Arctic are
(Shaw 1991b) has proved this hypothesis. Concentrations
much lower than the concentrations measured around major
of several heavy metals measured in the air over Greenland
point sources of emissions located mostly in the subarctic re-
(Heidam 1984) and at stations in Canada, such as Alert,
gion. For example, concentrations of Ni, Cu, and As mea-
Igloolik, and Mould Bay (Barrie and Hoff 1985, Barrie et al.
sured at several stations in northern Norway (Figure 7·26)
1989), are similar to those measured in the Norwegian Arc-
and on the Kola Peninsula are at least one order of magni-
tic and Alaska. The concentrations of V, Co, Zn, and Sb in
tude higher than the concentrations measured at Ny-Åle-
these parts of the Arctic are also in good agreement with the
sund (Figure 7·25). Maximum daily concentrations of these
concentrations measured at the stations in Severnaya Zem-
metals on the Kola Peninsula are at least two orders of mag-
lya and Wrangel Island in the Russian Arctic (Vinogradova
nitude higher than the concentrations measured during the
et al. 1993). A lack of data made the comparison for other
episodes of air pollution transport to Ny-Ålesund. Concen-
heavy metals impossible.
trations of several heavy metals near the Severonikel smelter
Winter air concentrations of several heavy metals in the
stacks on the Kola Peninsula are about three orders of mag-
Norwegian Arctic, Alaska, Greenland, the Canadian Arctic,
nitude higher than the maximum concentrations at Ny-Åle-
and Severnaya Zemlya and Wrangel Island in the Russian
sund (Panichev et al. 1993). Concentrations decrease by one
Arctic in the 1980s and the 1990s, together with European
order of magnitude within 20 km distance from the stacks.
emissions of As, Pb and Zn during the same periods, are
presented in Figure 7·25. Concentrations at the beginning of
7.6.1.3. Atmospheric deposition in the Arctic
the 1980s are somewhat higher than those at the beginning
of the 1990s. This is particularly true for Pb due to the re-
Although there have been a number of successful studies di-
duction in the use of leaded gasoline in various regions of
rected at the origin of Arctic air pollution by heavy metals, a
the Northern Hemisphere. However, the available data can-
quantitative assessment of the portion of the pollution load
not provide enough information to assess quantitatively the
actually deposited in the High Arctic has not been made. In-
decrease in concentrations from the 1980s to 1990s.
formation on deposition of heavy metals in the Arctic can be
The concentrations of heavy metals in Arctic air in winter
obtained from studies by Davidson (1991) who confirmed
are generally one order of magnitude higher than the winter
that atmospheric chemistry over the Greenland Ice Sheet dif-
concentrations measured in Antarctic air. Summer concen-
fers greatly from the patterns observed at coastal sites. This
trations in the Arctic and Antarctica are similar.
observation is very important when assessing the magnitude
Chapter 7 · Heavy Metals
403
Kirkenes
Karpdalen
ng/m3
Maximum
ng/m3
Average
ng/m3
Maximum
ng/m3
Average
150
15
150
15
100
10
100
10
50
5
Norway
50
5
0
0
0
0
Kirkenes
Ni
Cu As
Ni
Cu As
Ni
Cu As
Ni
Cu
As
Karpdalen
Holmfoss
Viskjøfjell
Viskjøfjell
ng/m3
Maximum
ng/m3
Average
ng/m3
Maximum
ng/m3
Average
150
15
150
15
Holmfoss
Pechenga
100
10
100
10
Finland
50
5
50
5
Svanvik
0
0
0
0
Ni
Cu As
Ni
Cu
As
Zapolyarnyy
Ni
Cu As
Ni
Cu
As
Kobbfoss
Nikel
Kobbfoss
Svanvik
ng/m3
Maximum
ng/m3
Average
Russia
ng/m3
Maximum
ng/m3
Average
150
15
150
15
100
10
100
10
Noatun
Study area
50
5
50
5
0
0
0
0
Ni
Cu
As
Ni
Cu As
Ni
Cu As
Ni
Cu As
Noatun
Birkenes
0
10
20 km
ng/m3
Maximum
ng/m3
Average
ng/m3
Maximum
ng/m3
Average
150
15
150
15
100
10
100
10
Birkenes
50
5
50
5
0
0
0
0
Ni
Cu As
Ni
Cu
As
Ni
Cu As
Ni
Cu
As
Figure 7·26. Average (15-month) and 24-h average maximum air concentrations of Ni, Cu and As measured at several stations in northern Norway
near to major point sources on the Kola Peninsula. (Source of data: Norwegian Institute for Air Research).
of atmospheric deposition of heavy metals on the basis of
1994). For Zn, Cd, and Cu, the estimated fluxes were about
measurements.
2500, 60, and 200 tonnes, respectively. There has been a
The past and present atmospheric deposition fluxes of
reduction in the use Pb in gasoline over the last two de-
heavy metals in the Arctic can be estimated by combining
cades. A corresponding decrease can be monitored in the
the heavy metal concentrations measured in the snow or
Greenland Ice Sheet (e.g., Boutron et al. 1991, 1995, Hong
ice with the snow/ice accumulation rates. Candelone et al.
et al. 1994).
(1996) have made these calculations for Summit in Green-
In Norway, surveys of Pb concentrations in moss have
land for different times: 7760 years bp, which gives the
been repeated since 1975 to estimate the latitudinal gradient
pre-human activity Holocene fluxes; 1773, which corre-
in Pb deposition (Figure 7·27, next page). These studies also
sponds to the onset of the first Industrial Revolution; 1850,
indicate a decreasing trend in Pb deposition (Steinnes et. al.
which is the time when concentrations of some heavy met-
1994, Steinnes 1995).
als started to increase; and 1992. The results of flux esti-
Other deposition measurements have also been carried
mates for Pb, Zn, Cd, and Cu at Summit are presented in
out in subarctic areas, such as the Kola Peninsula, northern
Table 7·8 for these periods, as well as for the 1960s and
Scandinavia, and Canada. The results of measurements on
1970s when the metal concentrations reached maximum
the Kola Peninsula carried out by Sivertsen et al. (1992) and
values. The authors then estimated the cumulative anthro-
Jaffe et al. (1995) as well as those summarized by Alexeyev
pogenic deposition flux of heavy metals to the Greenland
(1993) are in good agreement, suggesting that the total an-
ice cap from 1773 to 1992, i.e., from the onset of the first
nual deposition of Cu and Ni can reach a few hundred
Industrial Revolution to present. The cumulative deposi-
mg/m2 in the direct vicinity of the smelter stacks and de-
tion flux during this period was 3200 tonnes for Pb, which
crease to a few mg/m2 within a few tens of kilometers. An
is one order of magnitude higher than the deposition of the
example of Ni deposition on snow in the Kola Peninsula in
metal calculated for Greco-Roman times (Hong et al.
April 1990 is presented in Figure 7·28 (next spread) (Sivert-
sen et al. 1992). Wet deposition is far greater than dry depo-
Table 7·8. Heavy metal deposition fluxes at Summit, central Greenland.
sition in the area.
(After Candelone et al. 1996).
Rainwater composition was measured at eight Arctic
catchment areas in northern Europe: four in Russia, three in
Flux, ng/cm2/y
Finland, and one in Norway (Reimann et al. 1997). Close to
Year
Cd
Cu
Pb
Zn
industrial sources in Russia, most of the heavy metals mea-
7760 bp
0.006
0.039
0.013
0.53
sured (e.g., Co, As, Cu, Mo, Ni, Sb) show enrichment of
1773
0.006
0.064 a
0.18
0.37
two to three orders of magnitude in their median levels,
1850
0.006
0.053
0.35
0.7
1992
0.018
0.17
0.39
1.2
compared with background levels measured in Finland.
1960s/1970s maximum
0.041
0.22
2.5
0.041
The 1980s deposition of heavy metals on the Kola Pen-
insula was at least one order of magnitude higher than the
a. No satisfying plateau of concentration was obtained for Cu in the 1773
sample.
deposition in the 1960s (in Kryuchkov and Makarova 1989),
404
AMAP Assessment Report
1975
1985
1990
1995
2
5
10 20 30
40 60
80 100 120 140 Pb, mg/kg dw
Figure 7·27. Latitudinal gradient of deposition of Pb in Norway in different years, as reflected by Pb concentrations in moss. (After Steinnes et al. 1994,
Steinnes 1995).
with the major change occurring in the 1970s. The deposi-
area of the smelter stacks (Figure 7·28). Deposition of heavy
tion trend over the last 30 years in the region reflects the
metals in northern Scandinavia, also affected by emissions
emission trend discussed earlier in this chapter.
from smelters on the Kola Peninsula, is similar to the deposi-
The levels of annual deposition of Ni, Cu, and other
tion in southern Scandinavia resulting mostly from the long-
heavy metals in northern Finland, reported by Selin et al.
range transport of pollution from emission sources in Europe
(1992), are at the level of a few mg/m2 and agree well with
(Juntto 1992). Lead seems to be an exception, showing higher
measurements on the Kola Peninsula outside the immediate
deposition in the south.
Chapter 7 · Heavy Metals
405
dent in the water and sediments of aquatic environments
within 30 kilometers of the smelters. Soil samples at Kara-
sjok, Norway are clearly and heavily contaminated by Cu
(300-7400 g/g). The overall effects attributable to the
Kirkenes
Norway
smelters, constructed approximately 50 years ago, are de-
vastating. The cause is a combination of gaseous emissions
causing acid rain and a huge increment to the aerial deposi-
Karpdalen
tion of metals. Kryuchkov (1991) notes `industrial deserts',
Viskjøfjell
entirely or almost entirely void of vegetation, surrounding
Norway
the smelter towns for hundreds of square kilometers. Ac-
Holmfoss
cording to the 1989 State of the Environment Report for
Russia (State Committee for the Protection of Nature 1991)
the Kola Peninsula qualified as one of eight most seriously
Svanvik
polluted eco-catastrophe areas of the former Soviet Union.
Låg et al. (1970) have described the situation at Karasjok
as an example of natural heavy-metal-poisoned soil (and
Kobbfoss
Nikel
Zapolyarnyy
vegetation). This phenomenon is thought to be fairly com-
mon in areas where sulfide mineralization occurs in bedrock.
Russia
The `poisoning' results when heavy metals are extracted
from crystalline mineralization in bedrock and overburden,
then transported downslope in groundwater solution, and
finally reprecipitated in humus-rich soils where the solutions
containing heavy metals emerge at the surface (Bølviken et
al. 1977). Areas affected by this phenomenon often led to a
0
20 km
patchwork pattern where barren soil patches range in size
from 102 to 103 m2. Bølviken et al. (1977) have shown that
Ni deposition, mg/m2
such poisoned areas can be mapped using the LANDSAT
1
2
5
10
50
100
multispectral scanner system.
Metal concentrations in Greenland soils from four remote
locations were: < 12-37 g/g, < 12-13.8 g/g, 8-13 g/g,
Figure 7·28. Nickel deposition to snow on the Kola Peninsula in April
< 0.04-0.10 g/g, and < 0.01-0.03 g/g for Cu, Zn, Pb, Cd,
1990. (After Sivertsen et al. 1992).
and Hg, respectively. These concentrations probably repre-
sent background conditions. Results for Russian soils in
Yamal, Taimyr, and the Lena Reserve appear much lower by
comparison. Part of the explanation for this difference may
7.6.2. Terrestrial ecosystems
be natural. Another factor is analytical bias: results of the
The data upon which this section is based are provided in
joint Norwegian-Russian laboratory intercalibration (Akva-
Annex Tables 7·A1-7·A4.
plan-niva 1996) suggest that the Russian metal data could
be low by 30-50%. Notwithstanding analytical variability,
7.6.2.1. Soil
Somewhat limited data for metals in soils are available for
Greenland, Norway, Sweden, and Russia. General informa-
tion on concentrations of Pb, Cu, Cr, Ni, V, and Zn in the
soils of northern Fennoscandia are summarized in the Geo-
chemical Atlas of Northern Fennoscandia (Geological Sur-
veys of Finland, Norway and Sweden 1986). Figure 7·29
shows an example of available information for Norway. Ano-
malously high concentrations of these metals are noted; these
correlate with known geological provinces. The range of nat-
ural concentrations can be up to two orders of magnitude.
Large amounts of data have been collected in the Kola
Peninsula and in neighboring areas of Fennoscandia in an
attempt to quantify the effects of emissions from the metal-
lurgical complexes in the area (Pechenganikel in Nikel and
Zapolyarnyy, and Severonikel in Monchegorsk) and to study
the mobility of metals in soils in areas influenced by acidifi-
cation. The principal contaminants studied are Cu and Ni.
Barkan et al. (1993) and Evdokimova and Mozgova (1993a,
1993b) measured Cu in various soil horizons in areas up to
six kilometers from the Severonikel smelter over the period
1976 and 1986. Concentrations were highly variable (30-
4262 g/g) and indicate that even in an area receiving exceed-
ingly high loadings, some soil samples reflected near-back-
10
20
40 60
80 100 120 140 Pb, mg/kg dw
ground concentrations.
Contamination by both metals is clearly detectable through-
out the Kola region and, as already noted, is particularly evi-
Figure 7·29. Lead concentrations in soil in Norway in 1975.
406
AMAP Assessment Report
Table 7·9. Comparison of concentrations of metals in Arctic soils with soil quality guideline values. `Most stringent guideline value' is derived from Annex
Table 7·A17.
Most stringent
guideline value,
Metal
µg/g dw
USA (Alaska)
Canada
Greenland
Iceland
Norway
Sweden
Finland
Russia
Pb
25
n.d.
+
+
n.d.
+
n.d.
n.d.
n.d.
Cd
0.35
n.d.
++
+
n.d.
+++
n.d.
n.d.
++
Hg
0.1
n.d.
n.d.
+
n.d.
n.d.
n.d.
n.d.
+++
Se
0.7
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
+++ Virtually all reported values exceed the guideline. ++ Some (up to 50%) reported values exceed the guideline. + No reported values exceed the guideline.
it is probably valid to compare the 1994 and 1995 Russian
7.6.2.2. Microorganisms
data as they likely comprise an internally consistent dataset.
These data suggest no change in the metal content of soils in
No data were found for the metal content of microorgan-
Taimyr and Yamal as the differences in the values from 1994
isms in the Arctic terrestrial ecosystem.
to 1995 are generally less than the range of analytical error.
When the available circumpolar data are compared with the
7.6.2.3. Vegetation
most stringent soil quality guideline values for metals in soil
(Table 7·9) (BKH Consulting Engineers 1995), the frequency
The most abundant data on the metal content of vegetation
of exceedence occurs in the order Cd > Hg > Pb.
in the Arctic exists for mosses and lichens, particularly in
The background composition of soils is affected primarily
Scandinavia.
by bedrock geology, the hydraulic environment, weathering
Much of the data on lichens has been obtained in studies
processes, and biological processes. Deposited metals are
of the lichen reindeerwolf food chain. Data on the metal
distributed through the soil-water systems and trapped in
content of mosses are abundant because mosses have been
the soil profile with varying retention times according to
identified as useful monitors of atmospheric deposition of
their physico-chemical properties. The mobility of metals in
metals, Figure 7·27. The technique was introduced in 1968
the soil-water system is the result of a complex interaction
by Rühling and Tyler (1968) and is now being used routine-
of many processes. For many metals such as Al, Cd, and Zn,
ly for large-scale deposition studies in several countries. The
acidification of soils and waters is by far the most important
basis of the moss monitoring technique is that mosses lack a
factor regulating the concentrations and transport of metals
root system and therefore depend on surface uptake of chem-
from soils to running waters (Johansson et al. 1991). In con-
ical substances. They satisfy almost all their water and nutri-
trast to most other metals, the solubility of Hg in soils and
ent requirements directly from the air, and translocate them
waters does not increase with decreasing pH. The adsorp-
to their thin leaves. A dense moss carpet accumulates nearly
tion of Hg on humic matter actually increases at lower pH
all the mineral material deposited (Rühling et al. 1992).
values (Lodenius 1987).
A limitation of the technique is that factors other than air
The bioavailability of metals is related to chemical spe-
pollution contribute significantly to the element distribution
ciation. The divalent form of metals is usually the most
observed in mosses. For instance, windblown soil dust de-
readily bioaccumulated because of its mobility across cell
posited on the moss surface can be exceedingly difficult to re-
membranes. Consequently any factor which favors the en-
move prior to analysis (Steinnes and Jacobsen 1994). This can
hancement of divalent metal cations in the terrestrial/aquatic
lead to a relatively large sampling variability and ultimately to
system will lead to greater bioaccumulation. In the Arctic,
highly variable results. These problems pose the greatest diffi-
the most important factor in this regard in soil systems is
culty for interpreting and comparing results for metal loading
acidification (pH). Other important factors include redox
with various receiving environments. It is not known to what
potential, presence of complexing ligands and competing
extent surface contamination of mosses affects each dataset
ions, and temperature.
compiled in Annex Table 7·A2. In Table 7·10, the content of
Table 7·10. Metals in samples of the feather moss (Hylocomium splendens) collected in various regions of the Arctic.
Region
Year
Pb, µg/g dw
n
Cd, µg/g dw
n
Hg, µg/g dw
n
Se, µg/g dw
n
USA
1990-1992
0.35-1.6
21
< 0.03-0.98<
21
< 0.016-0.112<
Greenland
1.2-9.9
40
0.09-0.49
39
0.08-0.17
21
0.29-15.7
24
Norway
Svalbard
2.0-9.6
5
0.12-0.52
0.30-9.97
5
North Norway
1977
15
0.37
North Norway
1985
11
0.17
0.43
Russia
0.6-3.6
6
0.02-0.06
6
< 0.01-0.05<
6
Table 7·11. Metals in samples of lichen collected in various regions of the Arctic.
Region
Species
Pb, µg/g dw
n
Cd, µg/g dw
n
Hg, µg/g dw
n
Se, µg/g dw
n
Canada
Bathurst
Unspecified
0.01-0.15
12
0.01-0.09
12
0.01-0.08
12
Cambridge Bay
Unspecified
0.05-0.41
3
0.06-0.24
3
0.05-0.15
3
Inuvik
Unspecified
n.d.-0.08
8
0.01-0.08
8
n.d.-0.04
8
Greenland
Cetraria nivalis
0.8-6.4
62
0.07-0.16
62
0.03-0.05
24
0.01-0.22
40
Northern Finland
Hypogymnia physodes
17-.2
154
0-.57
154
Russia
Taimyr
Cetraria islandica
0.57-1.5
2
0.15-0.18
2
0.03-0.05
2
Taimyr
Cetraria delisei
0.-80
1
0-.08
1
0.-02
1
Various
Lichen/Lichenophyta
0.9-2.7
7
0.03-1.7
7
< 0.01-0.05 < 7
Chapter 7 · Heavy Metals
407
Pb, Cd, Hg, and Se in the species Hylocomium splendens are
7.6.2.4. Terrestrial birds
compared for samples collected in Alaska, Greenland, Norway,
and Russia. Large ranges are inherent in the data and are in
There is only a limited database on metal concentrations in
part related to sampling variability for the reasons noted above.
terrestrial bird species in the Arctic. The absence of time se-
Values for Russian samples appear to be low. This may be
ries measurements on any given species and of any system-
due to an analytical bias to underrecovery as noted in the re-
atic spatial sampling generally precludes evaluation of Arctic-
sults of the joint Norwegian/Russian laboratory intercalibra-
wide spatial and temporal trends. In general, concentrations
tion (Akvaplan-niva 1996). Differences in concentrations be-
in all birds decrease in the order kidney > liver > muscle.
tween samples collected in 1977 and 1985 in the Norwegian
The highest metal concentrations occur in grouse and
monitoring study (Steinnes et al. 1994) suggest no substantial
ptarmigan in all areas, and the highest concentrations for
temporal trend. Mäkinen (1994) measured the metal content
these species occur in Canada (Annex Table 7·A3). The
of several moss species along two transects (one north-south,
high concentrations of Cd in willow ptarmigan (Lagopus
the other east-west) in northern Lapland and the Kola Penin-
lagopus) kidney tissue of individual birds in Canada (1020
sula. For each metal studied, the transects effectively mapped
g/g dw) and in Norway (121 g/g dw) are notable. These
the zone of influence of emissions from various smelters and
are among the highest values for Cd reported for wild birds.
mining operations. Elevated values were found at distances
Wren et al. (1994) examined the concentrations of Cd,
of more than 100 km from the smelters. The highest values
Cu, and Zn in willow ptarmigan from Arctic Norway and
close to Monchegorsk were 263 times higher for Cu and
found that Cd concentrations were consistently three to ten
1328 times higher for Ni than the levels found in mosses in
times higher in the kidneys of adults than of juvenile birds
Finnish Lapland (background). Ranges of metals in mosses
from the same location. Zinc concentrations tended to be
found along the transects were as follows: 0.1-0.95 g/g,
higher in adults as well but the difference was only about
1-70 g/g, 5-1360 g/g, 3-3100 g/g, 2-25 g/g, and 20-50
10-30%. There was no difference in Cu concentrations be-
g/g for Cd, Cr, Cu, Ni, Pb, and Zn, respectively.
tween adult and juvenile willow ptarmigan. Wren et al. con-
Metals in various lichen species are compared in Table
cluded that the bioaccumulated Cd was natural in origin.
7·11 for samples collected in Canada, Greenland, Finland,
RCMA (1996) provide summary data for Cd, Pb, and Hg
and Russia. The lowest values of Pb and Cd are reported
in muscle and liver tissues of carnivorous, herbivorous, and
for samples collected in Canada, and the highest for Finnish
omnivorous birds across Siberia. In muscle tissue, concentra-
moss. The Hg content of lichens was similar (generally 0.01-
tions for all metals decrease in the order carnivores > omni-
0.10) for all regions.
vores > herbivores; in liver tissues there is no clear pattern.
Ford and Vlasova (1995 in press) measured Pb and Cu con-
Some spatial trends are evident in the data. The muscle and
centrations in the forage lichen Cetraria cucullata in Arctic Alas-
liver tissues of herbivores contain more Pb in the east than
ka and in the Taimyr Peninsula region of Russia north and
in the west of Siberia. A similar trend occurs for Cd in the
south of the huge Cu-Ni smelter at Norilsk. The data indicate a
muscle and liver of omnivores, and for Cd and Pb in the liver
zone of influence of about 200 km around Norilsk where Cu
tissues of carnivores.
concentrations are approximately 1 to 2 orders of magnitude
Ptarmigan provide the most diverse dataset at the present
higher than those measured at locations 150 to 500 km north
time. The data summarized below (and in Table 7·12, and
of Norilsk. Lead concentrations in C. cucullata were also high-
Figure 7·30 next page) provide comparative values for Pb,
er to the south but the enrichment was only about a factor of
Cd, Hg, and Se in the willow ptarmigan and the rock ptar-
5-10 relative to the lichens from the northern sites. The data
migan (Lagopus mutus).
show that values for Pb and Cu are generally low in both Arc-
Willow ptarmigan tend to contain more Cd in kidney tis-
tic Alaska and the Taimyr Peninsula relative to industrialized lo-
sue than do rock ptarmigan. This observation has also been
cations like Norilsk, but that samples of C. cucullata from Prud-
reported in these birds in Norway (Mykelbust et al. 1993,
hoe Bay and along the Dalton Highway (the land transporta-
Pedersen and Myklebust 1993) and in Russia (RCMA 1996).
tion corridor between Fairbanks and the Prudhoe Bay oilfields)
Different accumulation patterns of metal uptake among dif-
Alaska contain concentrations of Pb and Cu similar to those
ferent bird species (and even among individuals of the same
found in the industrialized regions of Siberia and western Russia.
species) are usually caused by differences in geological, di-
Table 7·12. Concentrations of Pb, Cd, Hg, and Se in willow ptarmigan (Lagopus lagopus) and rock ptarmigan (Lagopus mutus) for samples collected in
the Arctic regions of the USA, Canada, Norway, and Russia. Figures in brackets shows number of samples. (Sources of data: see Annex Table 7·A3).
Concentration µg/g, dry weight (except Russia, wet weight)
Lead
Cadmium
Mercury
Selenium
Area
Species
Liver
Kidney Muscle
Liver
Kidney Muscle
Liver
Kidney
Muscle
Liver
Kidney Muscle
USA, Rock
<1-2
3-18
21-109
< 0.02
0.48-1.3
Alaska
ptarmigan
(18 )
(18 )
(18 )
(18)
(18)
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Canada, Willow 0.08-1.5
0.1-7
20-122 9-1020
< 0.05-0.18
< 0.05-0.3
0.9-1.8
0.3-4
Yukon ptarmigan
(9)
(9)
(9)
(9)
(9)
(9)
(9)
(9)
Territory Rock 0.08-2.8 0.9-176
7-54
28-887
< 0.05-0.21
0.07-0.23
< 0.05-1.4 1.5-4.9
ptarmigan
(4)
(4)
(4)
(4)
(4)
(4)
(4)
(4)
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Norway Willow
0.43-1.47 0.45-0.9
0.75-9.1 46-121
0.019-0.046 0.062-0.093
0.48-0.93
ptarmigan
(64)
(49)
(64)
(49)
(64)
(49)
(64)
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Sweden
Willow
0.37-0.8 0.73-2.1
8-13
56-126
0.2-0.4
0.08-0.1
ptarmigan
(9)
(9)
(9)
(9)
(9)
(9)
------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Russia
Willow
0.3
0.12
0.50
0.36
0.06
0.04
ptarmigan
(1)
(1)
(1)
(1)
(1)
(1)
Rock
< 0.05
< 0.05
0.05
0.05
0.04
0.02
ptarmigan
(1)
(1)
(1)
(1)
(1)
(1)
408
AMAP Assessment Report
Liver
Kidney
2864
147524
53
472
11
22
62
104
15
103
21
118
7
57
6
47
28
229
6
74
Cd
4
9
Cd
68
128
µg/g dw
µg/g dw
60
600
11
129
50
500
13
160
5
58
40
400
30
8
300
56
20
200
13
126
10
100
0
0
Liver
Kidney
0.97
0.80
1.33
0.46
0.55
0.62 0.110.10
0.15
0.19
0.54
0.20
0.38
0.69
0.57
0.32
0.22
1.14
No data
0.16
Pb
1.30
0.88
2.40
0.82
0.15
0.97
Pb
µg/g dw
µg/g dw
0.59
0.37
2.40
2.40
0.45
1.50
2.00
0.70
2.00
0.96
1.60
1.60
1.20
0.37
1.20
0.73
0.80
0.80
0.80
2.10
0.40
0.40
0
0
0.15
0.75
0.04
Liver
Kidney
0.03
0.11
0.07
0.12
0.04
0.13
0.05
0.15
0.12
0.06
0.03
0.13
0.18
0.09
0.05
0.08
0.30
Hg
0.03
0.08
0.03
0.04
0.04
0.10
Hg
µg/g dw
µg/g dw
0.12
0.60
0.04
0.22
0.10
0.02
0.50
0.02
0.07 0.10
0.08
0.40
0.06
0.02
0.30
0.08
0.04
0.20
0.04
0.10
0.02
0.10
0
0
Figure 7·30. Circumpolar distribution of Cd, Pb and Hg levels in liver and kidney tissue of willow ptarmigan (Lagopus lagopus).
Chapter 7 · Heavy Metals
409
Table 7·13. Cadmium in liver, kidney, and muscle of Rangifer tarandus from various locations in the Arctic. (Sources of data: see Annex Table 7·A4).
Concentration µg/g, dry weight (except Finnish Lapland and Russia, wet weight)
Liver
Kidney
Muscle
Location
Range
n
Mean
Range
n
Mean
Range
n
Mean
Canada
Finlayson Herd
3-64
49
11.4
51-882
50
121
Bonnet Plume Herd
1.1-8.3
21
3.2
7-108
21
38
Porcupine Herd
1.1-15.2
81
4.6
7-176
80
41
Tay Herd
1.7-19.6
20
9.6
44-473
20
121
Northwest Territories,
Bathurst Herd
0.05-2.7
20
1.96
5-26
20
9.68
Arviat
1.2-6.6
10
3.69
10-63
10
33.9
Southampton Island
0
12-45
10
18.8
Cape Dorset
0.7-4.7
10
2.24
4-24
10
14.1
Lake Harbour
1.1-7.9
10
3.88
5-58
10
32
Inuvik, NWT
1.1-11
10
5.83
6-89
10
42.7
Beverly, NWT
2.1-4.7
10
3.42
14-59
10
31
Cambridge, NWT
0.6-4
10
1.35
4-19
10
9.4
Taloyoak, NWT
0.68-1.5
10
1.06
5-12
10
7.4
Pond Inlet, NWT
0.3-1.6
10
0.98
10-19
10
14.5
Denmark/Greenland
0.32-2.33
13
0.001-0.009
14
Finnish Lapland
Southern Lapland
0.190-0.402 a
60
0.525-1.72 a
90
< 0.001-0.002 a
59
Western Lapland
0.233-0.758 a
60
0.650-4.62 a
90
< 0.001-0.003 a
62
Eastern Lapland
0.388-0.958 a
51
0.938-4.25 a
97
< 0.001-0.003 a
66
Northern Lapland
0.310-0.546 a
48
1.03-2.84 a
48
0.002-0.006 a
30
Norway
0.1-1.7
52
0.4
0.3-10
52
1.5
Russia
0.05-0.08 a
5
0.05-0.45 a
4
<0.02-0.10 a
8
USA
0
38-61
3
51.5
a. Range of means, µg/g ww.
etary, and physiological factors. Myklebust et al. (1993)
that the Cd burdens found in caribou liver or kidney consti-
measured Cd in the food of willow and rock ptarmigan. The
tute a health risk to the animals themselves. Elliot et al. (1992)
Cd content of items in the willow ptarmigan diet (willow
and Nicholson et al. (1983) report that significant kidney dys-
shrubs, Salix spp.; and birches, Betula pubescens, B. nana)
function occurs for most mammals and birds when Cd con-
was generally higher than that of the rock ptarmigan diet
centrations exceed 100-200 g/g ww (approximately 400-
(more crowberries, bilberry, and mountain avens), suggest-
800 g/g dw). Values have been reported at these levels for
ing diet as an important factor in explaining the difference
caribou kidney in the Arctic only infrequently.
in body burden between willow and rock ptarmigan.
Concentrations of Cd in caribou tissues decrease in the
Pedersen and Myklebust (1993) showed that there is no
order kidney > liver > muscle in all regions. There is also a
net accumulation of Cd in the kidney or liver of ptarmigan
strong positive correlation between Cd concentration in kid-
after their first winter, suggesting that once a threshold is
ney and age in all regions (Gamberg and Scheuhammer
attained, uptake and excretion of Cd are balanced. A major
1994, Rintala et al. 1995, Frøslie et al. 1986). The correla-
mechanism for excretion of Cd could be through shedding
tion is much weaker in the case of Cd in liver, and a leveling
of feathers during molting. (It is known, for example, that
off of Cd concentrations tends to occur in older animals
a significant excretion mechanism of Hg is through feather
(> 10 years). Gamberg and Scheuhammer (1994) noted also
molting (Braune and Gaskin 1987)). The evidence provided
that concentrations of Cd in liver and kidney of caribou
by Pedersen and Myklebust (1993) supports this. They note
tended to be higher in samples collected in spring than those
that ptarmigan molt 3-4 times per year and replace 100-120
collected in autumn. If a conversion factor from wet weight
g feathers which amounts to 20-25% of their body weight.
to dw is four to one, then the Canadian, Finnish Lapland,
Furthermore, in willow ptarmigan, the lowest levels of Cd
Norwegian, and Greenland data fall within similar ranges.
in juvenile birds occur during the spring/summer molt (2-3
The Russian data appear to be low by comparison. It may
months), whereas the build-up of Cd in the liver and kidneys
be that the Russian data are low due to the analytical differ-
is seen during winter when the ptarmigan have had the same
ences discussed above (Akvaplan-niva 1996). Frøslie et al.
plumage for eight months.
(1986) also report on Svalbard reindeer based on the unpub-
The Nordic Council of Ministers (1992) has proposed max-
lished data of Norheim et al. There, median Cd concentra-
imum Cd concentrations for the kidney, liver, and muscle tis-
tions in liver and kidney were 0.6 g/g ww (n = 55) and 3.3
sue of meat (pigs and cattle) for human consumption of 1 g/g,
g/g ww (n = 44), respectively.
0.5 g/g, and 0.05 g/g (ww), respectively. By this measure,
A summary of Cd concentrations in liver, kidney, and
most birds in Annex Table 7·A3 exceed the guidelines.
muscle tissue of various caribou herds is given in Table 7·13
(data extracted from Annex Table 7·A4). Data for the Can-
adian herds located in the Northwest Territories indicate
7.6.2.5. Mammals
that there are clear differences in levels of Cd in the tissues
The most comprehensive data set for metals in Arctic mam-
among herds. The variation in Cd concentrations in the ani-
mals exists for caribou/reindeer (Rangifer tarandus). To les-
mals is thought to be strongly linked to natural sources,
ser extents both spatially and temporally, measurements are
namely that local vegetation (e.g., Salix spp.) becomes en-
compiled for a variety of other mammals in Annex Table
riched with Cd when it grows in soils derived from Cd-rich
7·A4. The data set for Rangifer tarandus is large because
mineral zones known to exist in the NWT and the Yukon
this species is known to bioaccumulate high concentrations
(Muir et al. 1996). There is, however, no systematic geo-
of Cd in liver and kidney tissue and consumption of these
graphical pattern in these differences. The values for Cd in
tissues may constitute a human health risk (Crête et al. 1989).
kidney of caribou of the Yukon herds however, were higher
There is no published evidence, however, that demonstrates
and more variable than those of the NWT. In Norway, there
410
AMAP Assessment Report
ng/g ww
ng/g ww
600
Liver
60
Muscle
500
50
400
40
300
30
200
20
100
10
0
0
1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993
1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993
Cadmium
Lead
Mercury
Figure 7·31. Changes with time in concentrations of Cd, Pb and Hg in liver and muscle tissue of reindeer (Rangifer tarandus) sampled in three districts
in eastern Sweden. (Source of data: Mats Olsson pers. comm.).
is a striking (approx. 3-fold) increase in Cd burdens in the
ground conditions. Mercury is high in Canadian animals
kidneys of reindeer from Arctic Norway to southern Nor-
and is generally attributed to natural geological sources in
way (Frøslie et al. 1986). A similar spatial trend has been
the Canadian Shield. Partitioning of metals among the var-
noted in Sweden. It is consistent with the gradients for acid
ious tissues varies with metal. For example, Hg tends to de-
precipitation (Overrein et al. 1980), atmospheric fallout
crease in the same order as does Cd (kidney > liver > mus-
(Hanssen et al. 1980), and the concentration gradients of
cle), whereas Pb generally decreases in the order liver > kid-
Cd in Norwegian surface soils (Allen and Steinnes 1980).
ney > muscle.
In Russia, the concentration of Pb in the muscle and liver
Among the other terrestrial mammals tested, the greatest
of reindeer are generally higher in the east than in the west;
enrichment of metals, particularly for Cd, occurs in moose
there is no similar trend for Cd (Melnikov et al. 1996).
(Alces alces). Extremely high values (up to 1380 g/g dw)
The only substantial temporal data set examining metal
have been reported in the Yukon. In Norway, moose kidneys
concentrations in caribou is provided by the Swedish Envi-
and liver were also enriched with Cd (max. value 19 g/g
ronmental Monitoring Programme (SEMP 1995) through
ww (kidney), and 3 g/g ww (liver)) and showed the same
which samples of reindeer have been collected continuously
north-south gradient as did reindeer. Frøslie et al. (1985)
since the early-1980s in three districts along the Swedish
point out that the same geographical differences for liver Cd
easternmost mountain chain, encompassing the Saami vil-
concentrations are seen in lambs grazing in natural pastures.
lages of Gabna, Laevas, Girgas, Iran, Ran, Handölsdalen,
Among the smaller mammals, examination of the data in
and Mittådalen. Analysis of the accumulated data (see Fig-
Annex Table 7·A4 led to the following observations:
ure 7·31) indicates that during the past ten years, Cd, Pb,
1. Lead, Hg, and Se concentrations tend to be low compared
and Hg concentration in the muscle and liver of Swedish
with those reported for similar tissues in reindeer and moose.
reindeer show no significant log-linear or linear change. Fur-
2. Some high kidney Cd values occur for beaver (130 g/g),
thermore, SEMP (1995) calculated that the number of years
porcupine (326 g/g), snowshoe hare (Lepus americanus)
required to detect an annual change of 5% ranged between
(166 g/g), ground squirrel (538 g/g), and european or
10 and 21 years depending upon element and tissue.
mountain hare (Lepus timidus) (99 g/g). Most of the
As noted above, the Nordic Council of Ministers (1992)
high values occur in animals from the Yukon. Some high
has proposed a maximum concentration for Cd in kidney,
values are seen in hares in Norway. The high values prob-
liver, and muscle tissue of meat (pigs and cattle) for human
ably are related to the diet of the animals, most of which
consumption of 1 g/g, 0.5 g/g and 0.05 g/g (ww), re-
have feeding preferences for and a corresponding accessi-
spectively. According to this proposed guideline, all caribou
bility to vegetation that accumulates Cd.
in Canada would fail. Far fewer exceedances would occur
3. Data for the lichen reindeer wolf food chain indicate
for Finnish reindeer. Rintala et al. (1995) note that Cd con-
no biomagnification of Cd, Pd, Hg, or Se.
centrations in the muscle samples of reindeer from Lapland
4. The data for Russia again appear to be the lowest of all.
were at the same level as those in muscle samples of Finnish
The reason(s) may be related to analytical biases as noted
pigs and cattle. The same trend has been seen for Pb and Hg
above.
in muscle tissue of Finnish reindeer.
Lead and Se in caribou liver, kidney, and muscle tissues
The ecological significance of the high Cd concentrations
generally fall within the ranges considered to represent back-
measured in the kidney of various mammals is currently un-
Chapter 7 · Heavy Metals
411
Table 7·14. Occurrence of cadmium concentrations exceeding 100, 200,
600, and 800 µg/g (dw) in the kidney tissue of mammals and birds of
7.6.3. Freshwater ecosystems
selected regions of the Arctic. Y indicates Yukon.
7.6.3.1. Metals in freshwater
n
n
n
n
Max.
Data on the background metal content of freshwater in the
>100
>200
> 600
> 800 value
Arctic are very limited both temporally and geographically.
Species/area
n
µg/g
µg/g
µg/g
µg/g
µg/g
Erickson and Fowler (1987) report data for 1985 and 1986
Caribou
in various channels of the Mackenzie River Delta during
Yukon
winter. Mean values were approximately 1.1-1.4 g/L, 0.5-
Bonnet Plume Herd 22
1
0
0
0
108
Finlayson Herd
52
30
9
2
1
882
2.5 g/L, 0.15-0.99 g/L, 10-123 ng/L and 5-8 ng/L for Cu,
Porcupine Herd
76
3
0
0
0
176
Zn, Pb, Cd, and Hg, respectively. Average values published
Tay Herd
20
10
23
0
0
473
for the same metals during the period 1960-1974 (Environ-
Alaska
40 Mile Herd
3
0
0
0
0
61
ment Canada 1981) were 5, 8, 2, 1, and 0.005 g/L, respec-
NWT
tively. Rovinsky et al. (1995) report values for Russian rivers
Beverley Herd
10
0
0
0
0
59
in generally the same ranges as noted for the Mackenzie
Bathurst Herd
20
0
0
0
0
26
Arviat
10
0
0
0
0
63
River, namely 0.3-3 g/L, 1.8-2.6 g/L, 0.15-0.99 g/L,
Southampton Island 10
0
0
0
0
45
20-290 ng/L, and 20-66 ng/L for Cu, Zn, Pb, Cd, and Hg,
Cape Dorset
10
0
0
0
0
24
respectively. Recent data for 11 rivers in Arctic Canada in-
Lake Harbour
10
0
0
0
0
58
Inuvik
10
0
0
0
0
89
dicate average concentrations of < 1-2 g/L, < 1-2 g/L,
Cambridge Bay
10
0
0
0
0
19
< 0.7-1.3 g/L, < 0.2 g/L and < 0.2 g/L for Cu, Zn, Pb,
Taloyoak
10
0
0
0
0
12
Cd, and Hg, respectively (Jeffries and Carey 1994). Much
Pond Inlet
10
0
0
0
0
19
Norway
204
0
0
0
0
18
higher values are reported for lakes and ponds of the Arctic
Finland
325
0
0
0
0
15
National Wildlife Refuge, Alaska) (Snyder-Conn and Lubin-
Russia
5
0
0
0
0
4
ski 1993) (Annex Table 7·A7). The values are suspect be-
Wolf
Yukon
21
0
0
0
0
22
cause even the detection limits noted in the data usually ex-
NWT
30
0
0
0
0
8
ceed the maximum values measures for metals in other Arc-
Moose, Y
51
35
10
2
1
1380
tic freshwaters.
Lynx, Y
1
0
0
0
0
6
Red fox, Y
1
0
0
0
0
1
There are numerous examples of localized enrichment of
Marten, Y
7
0
0
0
0
8
metals in freshwater in the Arctic. Most of these enriched
Mink, Y
3
0
0
0
0
0.5
Hoary marmot, Y
1
0
0
0
0
46
areas are associated with mining operations. Garrow Lake,
Muskrat, Y
13
0
0
0
0
8
NWT is used as a tailings pond at a Pb-Zn mine in the
Short-tailed weasel, Y
8
0
0
0
0
7
Canadian Arctic and concentrations of Cu, Zn, Pb, Cd, and
Dall sheep, Y
4
0
0
0
0
30
Mountain goat, Y
4
0
0
0
0
16
Hg are typically in the order of 28 g/L, 360 g/L, 1.8
Porcupine, Y
6
4
1
0
0
326
g/L, 0.7 g/L and 0.03 g/L, respectively (INAC 1994). In
Snowshoe hare, Y
29
1
0
0
0
155
Arctic Russia and the Kola Peninsula, such mining opera-
Ground squirrel, Y
19
1
1
0
0
538
Red squirrel, Y
26
1
0
0
0
155
tions have caused enrichment of metals in the freshwater
Beaver, Y
14
5
4
0
0
256
ecosystem on a much larger scale. Values exceeding 10 g/L
Spruce grouse, Y
38
2
2
1
0
760
Cu and Ni generally occur to about a 30 km radius around
Ruffed grouse, Y
7
3
2
1
1
1020
Blue grouse, Y
3
0
0
0
0
11
metallurgical complexes in the Murmansk Region but effects
Willow ptarmigan, Y
9
7
5
4
4
1020
of these metals in aquatic systems is easily detected hundreds
Rock ptarmigan, Y
4
1
1
1
1
887
of kilometers from the source (ACOPS 1995). The highest-
White-tailed ptarmigan,Y
2
2
0
0
0
161
Misc. ptarmigan, Y
6
5
3
0
0
232
recorded concentrations of Cu and Ni in waters of the Mur-
Mallard duck, Y
2
0
0
0
0
4
mansk Region are summarized in Table 7·15. During the pe-
Scaup, Y
2
0
0
0
0
9
riod 1991-1994, Cu attained concentrations of 11-2524
times the maximal allowable concentration (MAC), and Ni
known. No clearly demonstrated effects on wildlife in the Arc-
2-135 times the MAC. The ecosystems of at least five water
tic related to Cd burdens have ever been reported. It is thought,
bodies are considered to be completely destroyed. Similar se-
however, that concentrations of Cd of 200 g/g ww (approx.
vere and widespread contamination of water bodies has
600 g/g dw) in kidney tissues of birds and mammals are
been reported for the Norilsk region, where intensive metal-
capable of causing renal dysfunction (Nicholson et al. 1983).
lurgical activities also occur.
Long-term exposure to lower concentrations is perhaps ca-
Data for metals in lakes of northern Scandinavia are gen-
pable of causing a number of sublethal effects.
erally within the lower end of the range reported here for
In Table 7·14, the data for Cd in kidney tissues are ar-
rivers (Annex Table 7·A7).
ranged to show the number of samples exceeding 100 g/g,
Table 7·15. Cu and Ni contamination of water bodies in Murmansk Re-
200 g/g, 600 g/g, and 800 g/g (dw) for all data in An-
gion (Hydrochemical Institute 1992, 1993, 1994, 1996).
nex Table 7·A4. The 100 g/g level identifies animals signifi-
cantly above circumpolar `background' concentrations; the
Highest recorded level, µg/La
600 g/g level identifies animals possibly at risk of kidney
Water Body
Metal
1991
1992
1993
1994
dysfunction; the 800 g/g level identifies animals having ex-
Kolos-Yoki River, mouth
Cu
47
14
29
27
treme body burdens of Cd.
Ni
102
60
195
53
The data clearly indicate that the greatest bioaccumula-
Luotn-Yoki River
Ni
56
38.5
32
17
tion of Cd in animals occurs in caribou, moose, squirrels,
Hayki-Lampi-Yoki River
Ni
32
43
24
24
grouse, and ptarmigan of the western Canadian Arctic (Yu-
Nyuduay River
Cu
2524
300
168
518
Ni
1347
409
465
400
kon). This is opposite to the trend noted for the marine eco-
Monche Lake
Cu
225
260
176
113
system (see section 7.4.4.8). It indicates that local natural
Imandra Lake (Monche-Guba) Cu
105
35
20
11
sources of Cd probably are the primary factor in controlling
Ni
195
6
37
5
bioaccumulation in the terrestrial environment of the Cana-
a. The maximum allowable concentrations (MAC) for these metals are:
dian Arctic.
MACCu = 1 µg/L; MACNi = 10 µg/L.
412
AMAP Assessment Report
Table 7·16. Comparison of concentrations of metals in Arctic freshwaters with water quality guideline values. `Most stringent guideline values' are from
Annex Table 7·A17.
Most stringent
guideline value,
Metal
µg/L
USA (Alaska)
Canada
Greenland
Iceland
Norway
Sweden
Finland
Russia
Pb
0.1 ++
+++
n.d.
n.d.
n.d.
n.d.
++
++
Cd
0.045
++
+
n.d.
n.d.
n.d.
n.d.
+
++
Hg
0.012 ++
+
n.d.
n.d.
n.d.
+
n.d.
+
Se
5 n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
+++ Virtually all reported values exceed the guideline.
+++ Some (up to 50%) reported values exceed the guideline.
+++ No reported values exceed the guideline.
When the available metal data are compared with the
dient and the enrichment of Pb and Cd in the surface sedi-
most stringent water quality guideline value for metals in
ment was only about a factor of 2.
freshwater (Table 7·16), the frequency of exceedence oc-
Verta et al. (1989) provide profiles for Pb, Zn, Cd, and
curs in the order Pb > Cd > Hg.
Hg in sediment cores from 18 small headwater drainage and
seepage lakes in Finland. The trends are in general agreement
with the data of Skotvold et al. (1996) and Johansson (1989),
7.6.3.2. Metals in freshwater sediments
namely a gradient decreasing south to north.
7.6.3.2.1. River and lake bottom sediments
Lockhart et al. (1995) provide data for Hg in dated cores
Data for metals in freshwater sediments are compiled in
from ten lakes in Arctic Canada. Recent fluxes of Hg varied
Annex Table 7·A5, and include samples from the USA,
from 5 g/m2/y to approximately 50 g/m2/y compared
Canada, Greenland, Norway, Finland, Sweden, and Russia.
with fluxes of 0.7 to 31 g/m2/y estimated from the deepest
The greatest quantity of data have been produced in Scan-
portions of the cores. The enrichment factors for Hg (ratio
dinavia, where comprehensive studies have been completed
of concentration of Hg at top of each core divided by that at
to characterize rates of metal accumulation in forest lakes,
the bottom) ranged from 1.1-7.0 (mean, 2.4). These values
to quantify the effects of long-range transport, and to de-
are consistent with data reported in temperate areas of
scribe the mobility of metals throughout forest ecosystems.
North America. Furthermore, Mannio (1996) provides data
Unfortunately, the number of samples analyzed and their
for sediments in northern Finnish lakes where surficial Hg
geographical distribution are insufficient to be used to eva-
concentrations are 40-180 ng/g and enrichment factors are
luate or describe spatial or temporal trends.
1.8-2.2. Similar concentrations and trends have been noted
Within Scandinavia, the data show similar trends. Con-
in northern Norway and Sweden (e.g., Rognerud et al. 1993).
centrations of Cu, Zn, Pb, Cd, and Hg can be highly vari-
Viewed as a whole, the data suggest a widespread and con-
able even in adjacent water bodies. An important factor af-
tinuing input of Hg into Arctic sediments. Indeed, the results
fecting metal concentrations is pH of the water body. In gen-
shown in Figure 7·32 for Finnish and Canadian sediment
eral, a decrease in pH results in a release of metals from se-
cores are remarkably similar despite the wide geographical
diments. In the case of Zn and Cd, it inhibits deposition be-
separation between the sampling areas.
cause retention times for these metals in the water column
This may be related to the rather unique properties of Hg
are increased. Mercury behaves differently from the other
as a metal. Specifically, it is highly mobile, particularly in the
metals. A decrease in pH usually decreases the solubility of
gas phase, and is capable of being re-emitted from sediment
Hg because the adsorption of Hg on humic matter appears
and water. Successive deposition/re-emission cycles com-
to increase at lower pH values. The Hg adsorbed on to hu-
bined with decreasing average temperatures at higher lati-
mic substances is not easily displaced.
tudes could lead to enrichment of Hg in Arctic regions.
Skotvold et al. (1996) studied heavy metals in sediments
Mannio (1996) measured the accumulation of Cu, Zn,
from 91 lakes in northern and Arctic regions of Norway.
Pb, Cd, and Hg in four lakes located in Arctic Finland. The
They concluded that Pb, Cd, Zn, and Hg deposition was pri-
surface layers of each core were generally enriched relative
marily from long-range air transport sources, whereas Cu
to deeper layers and the maximum values usually occurred
was the result of local sources. These conclusions were in
within the top 5 cm.
agreement with the data reported by Steinnes and Henriksen
Rognerud et al. (1993) provide data on historical trends
(1993). On the northern Norwegian mainland, gradients for
on accumulation rates of Cu, Cd, Co, Ni, Pb, and Zn for
Pb, Cd, and Hg decreased south to north, while Cu and Zn
two Norwegian lakes (Dalvatn and Durvatn) from cores col-
had the opposite trend. The range of mean values for Cu, Zn,
lected during a joint Norwegian Russian expedition in
Pb, Cd, and Hg were 25-110 g/g, 80-175 g/g, 17-85 g/g,
1992. The lakes are located downwind of Russian smelters.
0.45-0.75 g/g, and 0.06-0.20 g/g, respectively. On Spits-
Rognerud et al. (1993) showed that the sediment accumula-
bergen and Bear Island, the average concentrations of Pb (25-
tion rates for each lake were constant during the past 50-60
90 g/g) and Cd (0.5-1.3 g/g) were generally higher than
years. Consequently, any changes in the rate of accumula-
those found on the northern Norwegian mainland, but the
tion of individual metals should be attributable primarily to
mean concentrations of Hg (0.062-0.098 g/g) were lower.
changes in atmospheric deposition. The ratio between the
Johansson (1989) reported similar trends for metals mea-
accumulation rates of metals in the most recent settled sedi-
sured in 54 lakes in Sweden (14 were located within the
ment and a reference sediment from the 1920s prior to the
AMAP region). Gradients for Pb, Zn, Cu, and Cd all de-
start of smelting operations indicate that the accumulation
creased south to north. Concentrations in the surface layer
rates of Ni, Cu, and Pb have increased by factors of 2-4. The
were 3-25 g/g, 5-40 g/g, 20-65 g/g, and 0.07-1.3 g/g
increase in Ni and Cu are attributed to the smelters as the
for Pb, Cu, Zn, and Cd, respectively. In the south, vertical
deposition of these metals elsewhere in Norway are low
gradients of the metals along cores suggested enrichment
(Rognerud and Fjeld 1990, Sivertsen et al. 1991). The in-
factors of about 50, 7, 4, and 2 for Pb, Cd, Zn, and Cu, re-
crease in Pb accumulation rates are thought primarily to be
spectively. In the north, Zn and Cu showed no vertical gra-
related to the general increase of Pb in the atmosphere of the
Chapter 7 · Heavy Metals
413
Lakes in Northern Finland
Hg, µg/g dw
Hg, µg/g dw
0
0.05
0.10
0.15
0.20
0.25
0
0.02
0.04
0.06
0.08
0.10
0.12
0
0
1992
1988
1992
1991
1988
1975
1988
1988
1980
1982
1947
1975
1974
1915
1957
1976
1965
1889
1942
5
5
1971
1964
1955
1863
1917
1895
1944
1848
1956
1931
1837
1860
1945
1920
1934
1820
1827
1924
10
1909
1803
10
1893
1912
1879
1898
15
1872
15
1885
Nitsijärvi
Serriamjärvi
1863
69o18'N 28o06'E
1874
69o11'N 26o54'E
Lake 222
Pahtajärvi
1857
69o27'N 29o10'E
1866
68o10'N 24o00'E
20
20
Depth
Depth
cm
cm
Lakes in Northern Canada
Hg, µg/g dw
Hg, µg/g dw
0
0.02
0.04
0.06
0.08
0.10
0.12
0
0.02
0.04
0.06
0.08
0.10
0.12
0
0
1979
1986
1991
1966
1977
1950
1967
1972
1957
1954
5
1917
1953
1930
5
1919
1892
1903
1921
1885
10
1865
1900
1911
1863
1835
1853
10
1846
1864
15
1817
1804
1833
15
20
25
20
30
25
Hawk Lake
Fox Lake
63o38'N 90o42'W
61o14'N 135o28'W
35
Far Lake
Kusawa Lake
63o42'N 90o40'W
60o20'N 136o22'W
30
40
Depth
Depth
cm
cm
Figure 7·32. Concentrations of Hg in dated sediment cores from lakes in Arctic Finland and Canada. (Source of data: Lockhart et al. 1995, Northern
Canada; Mannio 1996, Northern Finland).
northern hemisphere (Norto and Kahl 1991), but a small
Russia influenced by metallurgical complexes such as those
contribution is related to the smelters (Hagen et al. 1991).
on the Kola Peninsula and around Norilsk. In these areas,
Rognerud et al. (1993) also conclude that there is no obvi-
the concentrations of Ni, Cu, Co, Cd, and Hg in the surface
ous contribution of atmospheric deposition of Cd and Zn
sediments of lakes located up to 40 kilometers from the
from the smelters.
source of pollution exceed background values by 10-380
Dahl-Hansen and Evenset (1995) measured relatively low
times (ACOPS 1995). As lake sediments are excellent stor-
concentrations of Cu (4.4-12 g/g), Zn (76-114 g/g), Hg
age reservoirs for metals, enhanced metal concentrations are
(0.026-0.092 g/g), Cd (0.11-0.89 g/g), and Pb (2.9-8.5
likely to exist in these areas for many decades.
g/g) in Lakes Nyulay, Kotyol, and Kapylty in Arctic Rus-
The metal content of cores from four locations in Green-
sia. Far greater pollution by metals is seen in areas of Arctic
land (Riget et al. 1997c), four lakes in the Yukon Territory,
414
AMAP Assessment Report
Table 7·17. Comparison of concentrations of metals in Arctic freshwater sediments with sediment quality guideline values. `Most stringent guideline val-
ues' are from Annex Table 7·A17.
Most stringent
guideline value,
Metal
µg/g dw
USA (Alaska)
Canada
Greenland
Iceland
Norway
Sweden
Finland
Russia
Pb
15 ++
++
++
n.d.
+++
++
+++
++
Cd
0.6 ++
++
+
n.d.
++
++
++
+++
Hg
0.15 n.d.
+
+
n.d.
++
++
++
+
Se
1 n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
+++ Virtually all reported values exceed the guideline.
+++ Some (up to 50%) reported values exceed t
+++ No reported values exceed the guideline.
Canada (Lockhart and Muir 1996), and Lake Kyusyur, Rus-
major Siberian rivers are characterized by low concentrations
sia (Rovinsky et al. 1995) all were similar and probably in-
of dissolved and particulate forms of heavy metals. Dai and
dicative of the range of natural background values (Annex
Martin (1995) showed that colloidal material (104 Dalton
Table 7·A5).
0.4 µm) made a significant contribution to the so-called `dis-
When the available metal data are compared with the
solved' fraction of heavy metals in the Ob and Yenisey rivers
most stringent sediment quality guideline value for metals
(20-50% Pb, 40-70% Cd and Ni, up to 70-80% Cu). The au-
in freshwater sediments (Table 7·17), the frequency of ex-
thors conclude that there is a fundamental role of colloidal
ceedence occurs in the order Pb > Cd > Hg.
fractions in determining the behavior of trace metals in the
estuarine zone as well as in the control of their net discharge
into the Kara Sea. A comparison of some dissolved and par-
7.6.3.2.2. Freshwater particulates
ticulate inputs by the Ob, Yenisey, and Lena rivers into the
The two main datasets on suspended particulates/suspen-
Kara and Laptev Seas is presented in Table 7·18. As data
sions in freshwater systems (see Annex Table 7·A6) are
concerning seasonal changes in concentrations are absent,
provided by Erickson and Fowler (1987) (Mackenzie River)
the values in this table should be considered qualitative only.
and Melnikov (1991) (four Russian rivers). Values for the
A comparison of gross and net river fluxes of dissolved
Mackenzie River (44-77 g/g, 134-225 g/g, 17-32 g/g,
metals to the Eurasian Basin (Guieu et al. 1996) and of fluxes
0.57-0.68 g/g, and 0.090-0.210 g/g for Cu, Zn, Pb, Cd,
from the other sources to the Arctic Ocean (inflow with saline
and Hg, respectively) are somewhat higher than the values
waters, organic matter decomposition) are shown in Table
reported by Melnikov for the Mezen, North Dvina, Ob, and
7·19 (Gordeev and Tsirkunov in press). The data show that
Pechora Rivers (3-28 g/g, 19-50 g/g, and 4-29 g/g for
net river flux exceeds gross river flux 5-6 times for Cd, 2.5
Cu, Zn, and Pb, respectively). There is no obvious reason
times for Ni and 1.5 times for Cu mainly due to the desorp-
for the difference, but it is probably related to the particle
tion from particles in estuarine zones. Net and total fluxes of
size of the suspended fraction, its organic content, sample
Zn were approximately equal. The data also show that river-
size, and sampling methods. The data for the Russian rivers
ine input of heavy metals into the Eurasian sector of the Arc-
seem to be low in that the values are much lower than the
tic Ocean appears to be rather small compared with other
river sediments themselves. Given that the metal content
sources (Cd, 4%; Cu, 27%; Ni, 11%; Zn, 2%). Similar data
usually varies inversely with particle size and that suspended
for North America was not available for the assessment.
sediments are usually finer or no coarser than riverbed sedi-
Table 7·18. Riverine fluxes of dissolved and particulate heavy metals to the
ments, this is somewhat puzzling.
Arctic Ocean (t/y) a.
The anthropogenic contribution of metal concentrations
in the particles of riverine suspended matter can be evalu-
River
Cu
Pb
Zn
Ni
Cd
As
Hg Fe 103
ated by using the concept of an enrichment factor (EF) in
Ob
which the ratio of metals in suspended matter of a given
Dissolved
850
6
160
530
0.3
0.2
12
river is compared with the ratio of the same metals in sus-
Particulate
840
260
1700
630
3.3
0.8
940
Yenisey
pended matter of global rivers:
Dissolved
1000
4
820
340
1.0
0.2
10
Particulate
650
180
1300
450
13.6
0.3
320
(Me/Sr)susp
Lena
EF =
Dissolved
300
9
180
160
2.8
80
0.4
12
(Me/Sr)susp.global
Particulate
490
400
2500
550
4.6
4.0
590
The elements Sr, Al, or Li, which exist almost entirely as a
a. Ob and Yenisey: Dissolved metals (except Hg) Dai and Martin 1995,
Kravtsov et al. 1994; particulate metals (except Hg) Gordeev et al. 1995.
part of mineral crystal structure, are taken as reference ele-
Lena: Dissolved and particulate metals (except Hg) Martin et al. 1993.
ments (Thomas and Martin 1982). By using this index and a
Dissolved and particulate Hg in all rivers Cossa and Coguery 1993,
large number of samples, the waters of most Russian rivers
Coguery et al. 1995.
fall within global natural variability Gordeev and Tsirkunov
(in press).
Table 7·19. Gross river flux and net river flux of dissolved heavy metals to
the Eurasian (EA) Basin with other inputs to the Arctic Ocean (mol/y).
Other
Total
Riverine
7.6.3.2.3. River heavy metal fluxes
Gross river
Net river
fluxes
fluxes
input
flux to
flux to
to Arctic
to Arctic
as %
An assessment of heavy metal fluxes by Eurasian rivers to
Metal
EA Basina
EA Basinb
Ocean c
Oceanb
of total
the Arctic Ocean has been completed in recent years, to a
considerable extent as the result of the International Pro-
Cd
2.0 105
1.1 106
2.9 107
3.0 107
4
Cu
5.0 107
7.4 107
2.0 108
2.7 108
27
gramme `SPASIBA'. The research (Cossa and Coquery 1993,
Ni
1.6 107
4.4 107
3.6 108
4.0 108
11
Martin et al. 1993, Kravtsov et al. 1994, Coquery et al.
Zn
4.4 106
4.4 106
2.7 108
2.8 108
2
1995, Gordeev and Shevchenko 1995, Guieu et al. 1996,
a. Gordeev and Tsirkunov submitted; b. Guieu et al. 1996;
Rachold 1995) showed that downstream sections of the
c. Yeats and Westerlund 1991.
Chapter 7 · Heavy Metals
415
General trends in the data for Hg in fish tissue include the
7.6.3.3. Microorganisms
following:
No data were found for the metal content of microorgan-
1. Highest concentrations tend to be in predatory fish such
isms in the Arctic freshwater system.
as pike and perch (Perca fluviatilis).
2. There is a tendency for concentration to correlate posi-
7.6.3.4. Algae and plants
tively with length (age) for pike and perch and no correla-
tion with whitefish and Arctic char.
Very few data have been collected on the metal content of
3. The highest concentrations of Hg in fish are not necessar-
aquatic plant species in the Arctic. Iivonen et al. (1992) re-
ily correlated with the presence of high Hg concentrations
port on the Pb, Cd, Cu, Zn, and Ni content of two floating
in sediments. There are several reasons for this. First, Se is
water-plant species, Nuphar luteum, and Sparganium sp.,
known to reduce the uptake of Hg by fish (and biota in
from three lakes in Arctic Finland. Concentration mean val-
general) so the Se content of sediments must be factored
ues are summarized in Table 7·20.
into any model involving correlation. Second, the concen-
Table 7·20. Metal concentrations in two floating water-plant species,
tration of calcium and magnesium affect the bioaccumu-
Nuphar luteum, and Sparganium sp., from three lakes in Arctic Finland
lation of Hg by altering competition equilibria. Finally,
(µg/g dw), (Iivonen et al. 1992).
pH and buffer capacity have an impact on the bioaccu-
Species
Pb
Cd
Cu
Zn
Ni
mulation of Hg by fish because humic substances can form
strong complexes with Hg thereby effectively reducing Hg
Nuphar luteum
0.1, 0.2
0.06, 0.09 1.35, 2.75
2.4-31
0.3-0.7
Sparganium sp.
0.57-4.48
0.15-1.14
2.2-5.6
35-128
0.8-2.0
availability (bioaccumulation of metals other than Hg is
usually enhanced in acidified low calcareous lakes).
4. The fish from Arctic Canada generally have the highest
body burdens of Hg found in the Arctic. This is thought
7.6.3.5. Metals in freshwater invertebrates
to reflect the widespread naturally elevated baseline for
Hg in the Canadian Arctic and may be related to Hg-or-
Very few measurements have been made on the concentra-
ganic matter associations). Fish exceeding the guideline
tions of metals in the lower food chain, including freshwater
level of 0.5 g/g for human consumption occur most fre-
invertebrates. The few data that have been reported are
quently in Arctic Canada and Greenland (range of mean
summarized in Annex Table 7·A8. Most values for Pb, Cd,
muscle values 0.17-1.0 g/g (n = 86)).
and Hg are at or near the detection limits.
The bioaccumulation of metals in fish is highly variable in
7.6.3.6. Fish
the Arctic freshwater ecosystem, as it is in other ecosystems.
Metal concentrations in various freshwater fish are com-
The partitioning of metals among fish tissues varies with
piled in Annex Table 7·A9. In general, there are few in-
species, age, sex, and season. Lake or river chemistry, the
stances of comprehensive data for a given species that will
presence of other contaminants, humic substances and other
allow comparison of values across the Arctic or as a time se-
complexing agents, and feeding habits of the fish also play
ries. A possible exception is Hg, for which the largest set of
significant roles in the dynamics of bioaccumulation.
Table 7·21. Mercury concentrations in Arctic freshwater fish.
Mercury concentration, µg/g wet weight
Region
Arctic char (Salvelinus alpinus) N
Whitefish (Coregonus spp.) N
Burbot (Lota lota)
N
Northern Canada
0.01-0.57
226
0.01-2.49
102
0.11-0.30
4
Greenland
0.17-0.99
86
-
-
Finnish Lapland
0.09-0.32
55
0.06-0.28
18
0.-23
8
Iceland
0.02-0.03
2
-
-
Norway
0.03-0.25
39
n.d.-0.18
99
-
Russia
-0.01
2
0.01-0.11
60
0.-01
1
Sweden
-0.10
?
0.28
?-
-
values exists. This emphasis on Hg in the environment and
Hg
in the food chain reflects the fact that it clearly bioaccumu-
ng/g ww
lates and that the effects Hg has on fish and on human con-
500
sumers are well documented (see sections 7.5.3 and 12.2.3.1).
Data for Hg in the muscle tissue of Arctic char (Salve-
linus alpinus), whitefish (Coregonus spp.), and burbot (Lota
400
lota) are relatively consistent for all Arctic regions as shown
in Table 7·21 (data extracted from Annex Table 7·A9).
300
The recent analysis of archived samples of pike (Esox lu-
cius) muscle tissue in Sweden under the Swedish Environmen-
tal Monitoring Programme provides the most comprehensive
200
time series for Hg in Arctic freshwater fish. A total of 237
samples of pike caught in 21 different years between 1968 and
100
1996 were analyzed (Annex Table 7·A9). Geometric mean
values (weight adjusted) varied between 194 and 459 ng/g
0
ww. No clear temporal trend is evident in the data (see Fig-
1968 1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996
ure 7·33); variations around a mean for the entire dataset of
Figure 7·33. Changes with time in (weight adjusted, geometric mean) con-
306 ng/g ww probably represent natural year-to-year fluctua-
centrations of Hg in muscle tissue of pike (Esox lucius) in Sweden. (Source
tions and analytical variability (Olsson pers.comm.).
of data: Mats Olsson pers. comm.).
416
AMAP Assessment Report
7.6.3.7. Metals in aquatic birds
7.6.5. Marine ecosystem
Based upon a limited number of measurements (Annex
The aim of this section is, for the Arctic marine ecosystem,
Table 7·A3), the bioaccumulation of metals by aquatic
to compile baseline data on the toxic metals Hg, Pb, Cd, and
birds appears to be at least an order of magnitude less than
Se; to present a detailed summary of the current state of
that of terrestrial birds. The distribution among tissues is the
knowledge of heavy metal levels; and to use this information
same as for terrestrial birds, however: kidney > liver > mus-
to identify temporal and geographic trends. Reports of levels
cle. There are too few data to delineate any spatial, tempo-
of heavy metal contamination likely to have been affected by
ral, or species trends. The largest dataset for metals in aqua-
local anthropogenic sources (e.g., mining) were not included
tic birds is for a group of ten species at various locations in
in the tables of this review, but levels found at these point
Arctic Canada (NWT, Yukon, Quebec, and Labrador, CWS
sources are described in the text.
1996). For Cd, Pb, and Hg, many values were at or near the
A review of the state of knowledge of contaminants in the
detection limit. Ranges for values exceeding the detection
Arctic marine ecosystem (to early 1990) was presented by
limit in breast muscle tissue were as follows: Cd, 0.02-0.87
Muir et al. (1992). This review focused mainly on the North
g/g ww (n = 134); Hg, 0.2-1.93 g/g ww (n = 158); and Se,
American part of the Arctic as well as Greenland. No data
0.04-2.1 g/g ww (n = 418).
were presented from Iceland, Scandinavia, or the Russian
part of the Arctic, and no data were presented for marine
birds. The following section provides an update to the re-
7.6.3.8. Mammals
view by Muir et al. (1992), and adds data from recent publi-
The data for metals in freshwater mammals consist of Pb,
cations as well as data obtained during the AMAP monitor-
Cd, and Hg measurements in the hair of ringed seals living
ing program. Data from Finland and Sweden have not been
in Lake Saimaa, Finland (Hyvärinen and Sipilä 1984). Al-
included in the marine review as the Baltic and the Gulf of
though Lake Saimaa is located just outside the AMAP as-
Bothnia are not within the agreed boundary of the AMAP
sessment area, it is nonetheless relevant to include this
Arctic marine area.
dataset here because the boreal environment of this lake is
In preparing this updated review, a thorough survey of
similar to that for lakes in the AMAP area. Ringed seals ex-
the literature has been carried out. Literature published since
ists in the Arctic and the data provide one of the very few
1991 was included in a search profile for heavy metals in the
observations of possible links between exposure to metals
Arctic environment. Data tables presented in `Arctic marine
and biological effects.
ecosystem contamination' by Muir et al. (1992) have been
Hair samples collected around 1983 had the following
incorporated into this report, and provide the basis for much
ranges of concentration means (n =32): 0.5-0.7 g/g dw, 3.6-
of the literature published prior to 1991 for the Canadian
8.5 g/g dw, and 3.2-20.7 g/g dw for Cd, Pb, and Hg, re-
Arctic. The eight Arctic countries also provided some litera-
spectively. According to data obtained for museum samples,
ture for the AMAP process. National assessment reports
the concentration of Hg in the hair of adult ringed seals was
from Greenland and Canada, recently compiled, have been
about the same as found by Hyvärinen and Sapilä (1984) in
included (Dietz et al. 1996, 1997a, Riget et al. 1997a, Muir
yearlings. In 1965 the concentration in hair was approxi-
et al. 1996). Electronic data searches were made in the North-
mately 50 g/g. The authors speculated that the sharp de-
ern Aquatic Food Chain Contamination Database (Version
cline in the ringed seal population of Lake Saimaa during the
1.0 E; October 1993) and in the databases of the Department
1960s and 1970s could be related to the insufficient avail-
of Fisheries and Oceans (Canada) in Winnipeg.
ability of Se in the lake, which made the seals more suscepti-
All data extracted from available literature are presented
ble to the toxic effects of Hg and premature (still) births.
in tabular form using a format similar to that used by Muir
et al. (1992). Species are presented in systematic order; geo-
graphical areas are given from north to south; and data are
7.6.4. Wetland ecosystems
arranged by year. Different tissue compartments (where ap-
Wetlands are a transitional compartment between the ter-
plicable) are listed in a consistent order (liver, kidney, mus-
restrial and aquatic environments and play an important
cle, bone, blubber).
role in fate of contaminants, heavy metals in particular. In
It is important to note that the level of information avail-
areas of strong anthropogenic impact, wetlands can serve as
able for specific metals varies considerably between loca-
accumulators of heavy metals and as sources of significant
tions and species and is related largely to the ease of sam-
secondary contamination of river waters. The survey of
pling. Most samples are collected near shore or in areas of
more than 250 wetlands in the Russian sector of the Arctic
human activity. For example, there are more data available
(Annex Table 7·A16) has shown that concentrations of
for metals in seals than in whales because it is easier to col-
heavy metals in wetland ecosystems are generally low, with
lect seal samples.
the exception of the areas where intensive heavy industry
An attempt was made to standardize the data where pos-
occurs. Three areas with high levels of local heavy metal
sible. In the older literature, little emphasis was placed on
pollution of wetland ecosystems have been identified: statistical requirements such as normal distribution, homo-
1) Kola Peninsula in the vicinity of non-ferrous smelters,
geneity of variance, etc. In order to facilitate statistical com-
2) the Vorkuta area on the north of the Komi Republic, and
parison, therefore, data are given primarily as geometric
3) the Norilsk area in Central Siberia. In the same areas,
means (GM) and relative standard deviations (rel. SD). Un-
increased concentrations of heavy metals in river waters
like sediment and water data, heavy metal data for biota are
have also been observed. Metal concentrations in remote
usually not normally distributed.
parts of the same regions are close to background levels,
Where possible, all concentrations are related to wet weight
with the exception of Far North-East Asia where increased
of tissues; where the concentrations given in the literature
concentrations of some metals including Hg are observed.
were given on a dry weight basis, they were converted to a
This area is situated in a zone of mercury ore belts, and in-
wet weight basis, using the moisture content presented in
creased concentrations of Hg in this area are thought to
each report for individual samples. Where such information
originate primarily from natural sources.
was not available, appropriate conversion factors were used.
Chapter 7 · Heavy Metals
417
Where values for both arithmetic means (AM) and geo-
Pacific contain ten times more Pb now, and those of the Arc-
metric means (GM) were given, but not relative SD, only
tic even more, than in prehistoric times. In the Greenland Ice
AM and SD are given in the table. Where raw data were
Sheet, modern Pb concentrations are 300 times higher than
given, GM and rel. SD are calculated. Where two or fewer
those characteristic of prehistoric times (Ng and Patterson
samples were analyzed, no rel. SD is calculated. Where the
1981). However a recent decrease has been observed due to
given values were means of subsamples taken from the same
the use of unleaded gasoline (Boutron et al. 1991, 1995,
animal, no SD is given.
Hong et al. 1994).
Statistical analyzes involving tests of the significance of
There is clear evidence that mining in the Arctic has in-
parameters and of differences of metal levels have not been
creased local Pb concentrations in seawater. At the Black
performed for this assessment, but such analyzes have in
Angel Pb-Zn mine at Maarmorilik in northwest Greenland,
many cases been carried out in connection with previous
which operated from 1973 to 1990, Pb levels in seawater as
reporting and publishing of the data. References to these
high as 200 000 ng/L were reported in bottom fjord water,
analyzes are given in the text.
where mine tailings were discharged (Asmund 1992a, 1992b).
In surface waters close to the mine site, Pb concentrations
up to 42 000 ng/L were measured (Asmund et al. 1991).
7.6.5.1. Seawater
Close to a cryolite mine in Ivittuut in south Greenland
Information on the concentrations of metal levels in seawa-
(closed in 1986), Pb concentrations in surface waters
ter from the Arctic is available mostly through Canadian
reached 18 000 ng/L (Asmund et al. 1991). Lead concentra-
and German programs. Concentrations of as many as 29
tions in the range of 40-100 ng/L were reported by Thomas
elements are reported from various parts of the Arctic. Arc-
et al. (1984) for the waters of Strathcona Sound, a northern
tic data on baseline concentrations have been compiled at
Baffin Island fjord at the Nanisivik Pb-Zn mine; these con-
the Institute of Ocean Sciences, Department of Fisheries and
centrations were approximately one to two orders of magni-
Oceans in Canada for the Beaufort Sea (Thomas et al.
tude higher than open ocean background concentrations.
1990), Northwest Passage (Thomas et al. 1991), Queen Eliz-
abeth Islands (Thomas et al. 1986a), and Canada Basin-Arc-
Cadmium
tic Ocean (Thomas et al. 1986b).
Moore (1981) reported Cd concentrations of 70 ng/L for the
Data from the Barents Sea, the White Sea and the Arctic
Central Arctic Ocean for the upper 100 m and then 20 ng/L
seas have been presented by the Ministry for Environment
to a depth of 2500 m, remaining approximately constant
Protection and Natural Resources of the Russian Federation
over the entire interval. Campbell and Yeats (1982) measured
(MEPNR 1994). It was concluded that many pollutants, in-
Cd for the eastern Arctic Ocean; concentrations ranged from
cluding heavy metals, enter Arctic waters through river in-
13 to 21 ng/L at depths of 1500-2000 m, but surface values
put from north-flowing rivers such as the Pechora, Yenisey,
were lower (8 ng/L). In the Norwegian Sea, a Cd concentra-
Lena, Ob, and Kolyma. Most of the pollutants accumulate
tion of 22 ng/L was measured in surface water (Danielsson
in river mouths and estuaries and then are spread across the
et al. 1985). Rosgidromet (1995) reports Cd levels between
Arctic basin toward Alaska by the circumpolar current. Re-
20 and 260 ng/L for waters of the Pechora, Kara, and Lap-
cent data are available from Pechora, Kara, and Laptev Seas
tev Seas. A comparison of values for the North Atlantic and
(Rosgidromet 1995).
the Norwegian Sea shows no significant mid-oceanic differ-
Contaminants, including heavy metals in seawater, have
ences for Cd (Mart et al. 1984). In relatively warm surface
been given low priority in the AMAP programs, because the
waters, e.g., the northeast Atlantic and more southerly lati-
concentrations of most pollutants in seawater are close to
tudes, Cd concentrations are lower in surface water than in
the detection limits of most laboratories and are very costly
deep water, indicating uptake of Cd by organisms at the sur-
to measure.
face, and release of Cd from sinking organic matter in sub-
surface layers.
Lead
Similar to the case for Pb, elevated Cd concentrations
Mart and Nurnberg (1984) reported Pb concentrations in
have been observed locally as a result of mining activities.
eastern Arctic ocean water of 15 ng/L at the surface and 3-4
In bottom water at the point of tailings discharge from the
ng/L at 1500-2000 m. Background concentrations for Pb in
Black Angel mine in Greenland, Cd concentrations of up to
Beaufort Sea shelf waters appeared to be in the range of
2500 ng/L were found (Asmund 1992a, 1992b). At the Na-
< 20-40 ng/L (Thomas et al. 1982). Recent seawater ana-
nisivik mine on Strathcona Sound in the Eastern Canadian
lyzes from the Pechora, Kara, and Laptev Seas in Russia
Arctic, Thomas et al. (1984) reported concentrations in sur-
show very high Pb levels in the range 160-500 ng/L (Ros-
face water from 30 to 130 ng/L. Close to Ivittuut in south
gidromet 1995), but these data should be confirmed as they
Greenland, surface water concentrations of Cd were between
may be unrepresentative. By comparison, Pb levels ranged
18 and 252 ng/L (Johansen et al. 1995).
from 29 to 41 ng/L in the surface waters of the North At-
lantic and the Norwegian Sea (Mart and Nurnberg 1984).
Mercury
In the North Pacific between Hawaii and California, Pb in
Very little information is available on the concentrations of
surface water ranged from 5 to 15 ng/L, decreasing with
Hg in Arctic Ocean water. Weiss et al. (1974) reported mean
depth to approximately 1 ng/L (Schaule and Patterson
Hg concentrations of 11-22 ng/L (0-400 m) for the southern
1981). Lead concentrations are consistently higher in sur-
Beaufort Sea. Thomas (1983) reported background values of
face water than in deeper layers. In coastal waters, concen-
1-15 ng/L for total Hg in samples from the Beaufort Sea shelf.
trations can reach 50 ng/L, and can be much higher in heav-
The dissolved Hg concentrations were strongly and nega-
ily polluted waters (Burnett et al. 1980).
tively correlated with the concentrations of particulate or-
Probably more than any other metal, Pb has been en-
ganic carbon. In Puget Sound (Washington, USA), the Hg
riched, particularly in the northern hemisphere, because of
concentration was 2-10 ng/L (Bothner and Robertson 1975).
anthropogenic inputs. An indirect estimate of 0.6 ng/L of
By comparison, the average Hg concentration was 4.1 ng/L
Pb in prehistoric oceanic surface water has been reported
in the Gulf Stream (over a depth of 250-4500 m), 8 ng/L (0 -
(Schaule and Patterson 1981). Surface waters of the North
750 m) in the Sargasso Sea, and 3-4 ng/L in waters around
418
AMAP Assessment Report
the United Kingdom (Mukherji and Kester 1979). In the
As
temperate zone in the North Pacific, total Hg was 14 ng/L
in surface waters and 2 ng/L between 500 and 5000 m
(Miyake and Suzuki 1983). Off the Swedish west coast, the
values ranged from 5 to 12 ng/L (Gustavsson and Edin
1985). There appears to be no great difference in the con-
centration of Hg in oceanic water around the world, except
in polluted waters where concentration can exceed the val-
ues noted above by an order of magnitude or more.
The older data have, however, been questioned in a re-
cent NATO workshop (NATO-ARW 1996). Current con-
sensus is that total dissolved Hg in unpolluted ocean water
is between 0.2 and 1 ng/L with somewhat higher concen-
trations in coastal waters and in density interfaces where
accumulation due to particulate dissolution is enhanced
(up to 2 ng/L). Methylated Hg (when it can be measured)
is generally less than 10% of the total Hg. Dimethylmer-
cury is often the major methylated species.
As, ng/g
7.6.5.2. Sediments
< 25
Concentrations of trace elements in sediments depend on
25-50
ng/g
local geology, particle size, the amount of organic matter,
50-75
150
and the degree of various kinds of anthropogenic influence.
75-100
The relationship between sediment metal concentration and
100
> 100
grain size, with higher metal concentrations found in fine
50
grained sediments than in sandy sediments, is a common
Samples with
non-fine grain
coarse or unknown
feature related to the geochemical distribution of total met-
samples
sediment texture
0
als with particle size. The anthropogenic influence might be
Cd
Pb
mg/kg dw
Year 0
10
20
30
40
50
60
70
80
2000
1980
1.330
0.660
1960
1940
1920
3.000
1900
St. 1
St. 3
St. 9
1880
Background
1860
0.650
Figure 7·34. Lead in sediments near the cryolite mine at Ivittuut, south
0.630
0.740
Greenland (after Johansen et al. 1995). Stations 1, 3 and 9 are located at
distances of 800 m, 1600 m and 5600 m, respectively, from the mine.
Cd, ng/g
mg/kg dw
< 0.1
0
100
200
300
400
500
600
800
0.1-0.2
ng/g
0
3
0.2-0.4
2
0.4-0.6
2
> 0.6
4
1
Samples with
non-fine grain
6
coarse or unknown
samples
sediment texture
0
8
Figure 7·36. Continued on next page
10
Concentrations of metals in fine-grained surface marine sediments in the
Arctic. (Sources of data: see Annex Table 7·A10).
12
Zn
Pb
estimated by the use of normalizers, such as percent organic
14
matter, Al (Windom 1989), percent grain size less than 63
16
microns (Allan 1971), other grain size fractions (Loring
Sediment depth
1990), and Li (Loring 1990). Although it is generally agreed
cm
that some kind of normalization is necessary when trace
Figure 7·35. Lead and zinc in a sediment core sampled in 1991, 1.5 km
element concentrations in sediments from different environ-
from the Black Angel mine in West Greenland. (Source of data: G. Asmund
pers. comm.).
ments are compared, or where anthropogenic influence is
Chapter 7 · Heavy Metals
419
Cu
Pb
88
Cu, ng/g
Pb, ng/g
< 20
< 7.5
ng/g
20- 40
7.5-15
ng/g
150
40-60
15-22.5
100
80
60-80
100
22.5-30
60
> 80
> 30
50
40
Samples with
Samples with
non-fine grain
coarse or unknown
non-fine grain
20
coarse or unknown
samples
sediment texture
0
samples
sediment texture
0
Hg
Zn
0.125
0.140
0.146
1.600
ng/g
Hg, ng/g
2.0
Zn, ng/g
< 0.02
< 50
1.5
ng/g
0.02-0.04
50-100
200
0.04-0.08
100-150
1.0
150
0.08-0.12
150-170
> 0.12
100
0.5
> 170
Samples with
Samples with
50
non-fine grain
coarse or unknown
non-fine grain
coarse or unknown
samples
sediment texture
0
samples
sediment texture
0
the subject of concern, the method of normalization that
The Norwegian data in Annex Table 7·A10 often repre-
should be used is still a matter of vigorous debate. Also ex-
sent sediments with significant anthropogenic influence.
cluded from the overview are severely polluted sediments
Wherever possible, concentrations are given both in surface
near local point sources, e.g., Maarmorilik (Loring and As-
sediments and in deeper layers expected to reflect natural
mund 1987), Strathcona Sound (Fallis 1982, Thomas and
background concentrations.
Erickson 1983), and Orkdalsfjorden (OSPARCOM/NIVA
Data for heavy metals in surficial Arctic marine sediments
unpubl. data). Other examples of contaminated sediments
are shown in Figure 7·36. Only those data for fine sediments
are given in Smith and Loring (1981). Some examples of the
(i.e. grain size reported as less than 63 microns or other in-
influence on sediments of mining in the Arctic are shown in
formation indicating this, cf. Annex Table 7·A10) are in-
Figures 7·34 and 7·35.
cluded in the plots in order to make the data comparable.
420
AMAP Assessment Report
On this basis, the geographical coverage is limited and there
ern Beaufort Sea continental shelf. The wide range of val-
is no systematic distribution evident in the data. The concen-
ues evident in the southern Beaufort Sea is probably related
trations of metals in the sediments correspond generally
primarily to sediment texture and mineralogy. The data for
with bedrock geology, and any anomalies can be explained
the Russian Arctic (7.7±1.5 to 17±5 mg/kg) could be low
in terms of known mineralization or anthropogenic activi-
by 0% and 61% based on the results of the Norwegian-
ties. If regional processes such as long-range air transport
Russian laboratory intercalibration exercise (Akvaplan-
were a primary factor in controlling sediment metal distribu-
niva 1996).
tion, then less variability would be expected in the data and
Loring and Asmund (1996) found that the concentrations
distribution, and an emission-related systematic pattern
of Pb in Greenlandic sediments from areas containing ter-
would have been observed.
tiary volcanic rocks and from the Nagssugtoqidian mobile
belt were lower (12-14 mg/kg) than in sediments from other
Copper
geological provinces (21-24 mg/kg). In contrast, Loring
The lowest and highest reported values of Cu are 4 mg/kg
(1984) found no dependence of concentrations on mud con-
in sandy sediments from the Canadian Arctic and 137 ng/kg
tent for a suite of Arctic sediments that contained 18±6 to
at Vaigat in West Greenland. The data reported for Arctic
21±8.5 mg/kg of Pb. The lack of enrichment of Pb in the
Russia are near the lower range of these values. These Rus-
upper one or two centimeters of Arctic marine sediments
sian results may, however, be low by 30-40% according to a
suggests that long range transport of aerosol particles is not
Norwegian-Russian laboratory intercalibration (Akvaplan-
an important process in the Arctic.
niva 1996). Copper concentrations in whole sediments from
Greenland are higher (90-140 mg/kg) in areas containing
Cadmium
tertiary volcanic rocks than in areas of non-volcanic sedi-
The lowest reported concentration of Cd in Arctic surface
ments (10 to 40 mg/kg) (Loring and Asmund 1996). The
sediments is 0.02 mg/kg in a sample from the Barents Sea.
concentration of Cu in Arctic marine sediments increases
High values of 4.6±0.3 and 1.33±0.54 have been reported in
with increasing content of fine grained sediment. Concen-
the southern Beaufort Sea and the Mackenzie Delta. These
trations can be as low as 4 mg/kg in sandy sediments. In
values probably include Cd from barite used in drilling flu-
sediments containing more than 95% mud, the Cu concen-
ids during oil and gas drilling activities (Thomas, pers. comm.).
tration in Arctic sediments was 61±18 mg/kg (Loring 1984).
Other high values include 0.32±0.31 for Beaufort Sea, 0.31
As the concentrations of Cu in recent (surficial) sediments
±0.08 for the MacCormick Fjord, Greenland, 0.23±0.21 for
are comparable to those in deeper and older sediments, the
the Ob Gulf, Russia, and 0.52 for one sample out of 43 in
levels summarized in Annex Table 7·A10 are believed to rep-
the Barents Sea. There are, however, several examples show-
resent true natural background levels.
ing that Cd accumulates at the interface between oxidized
and reduced sediment, where Mn also has a discontinuity
Zinc
(e.g., Gobeil and Macdonald unpubl.). For this reason, many
The lowest reported Zn concentration in sediments from the
sediment cores cannot be used for temporal trend studies.
Arctic is 9.5 mg/kg found in a sandy Barents Sea sediment.
No significant differences were found in Cd concentra-
The highest, 181±22 mg/kg, is reported for 50 samples from
tions between the geological provinces in Greenland (Loring
the Beaufort Continental Shelf. The data for marine sedi-
and Asmund 1996); this is probably related to the relatively
ments from the Russian Arctic, 25-75 mg/kg, could be up to
high analytical uncertainty in Cd measurements coupled
42% lower based on the results of the Norwegian-Russian
with the fact that most concentrations are low or at the de-
laboratory intercalibration (Akvaplan-niva 1996). Loring
tection limit. The average concentration reported for deep
and Asmund (1996) found no significant differences be-
water Greenlandic sediments was 0.l23±0.048 mg/kg. Lo-
tween Zn concentrations in sediments from various geologi-
ring (1984) found a strong dependence between Cd concen-
cal provinces of Greenland. The concentrations of Zn in fine-
trations and mud content; Cd concentrations ranged from
grained sediment ranged between 40 and 100 mg/kg.
0.058 to 0.144±0.030 mg/kg corresponding to mud content
Both Loring (1984) and Macdonald and Thomas (1991)
of < 5% to > 95%, respectively. There is no indication that
found that Zn concentrations were higher in fine grained se-
Cd accumulates in the upper layers of Arctic sediments.
diments than in sandy sediments. Zinc concentrations ranged
from 22 to 160 mg/kg in sediments containing 0% mud and
Mercury
100% mud, respectively. As there is generally no trend in the
Reported values for the concentration of Hg in Arctic sedi-
concentrations of Zn with depth, the values shown in Annex
ments range from below the detection limit of 0.01 mg/kg in
Table 7·A10 are believed to be true natural background lev-
subsurface sediments from Roroy and Skrova, Norway, and
els, with the exception of Norwegian surface sediments.
0.009±0.004 mg/kg in sediments from Uummannaq, Green-
land to 0.243±0.043 mg/kg in southern Beaufort Sea conti-
Lead
nental shelf sediments and 1.6±1.2 mg/kg in sediments from
The lowest reported levels are 1.4±0.2 mg/kg (Thomas et al.
Strathcona Sound, Canada collected in 1975. The latter very
1983, cited in Muir et al. 1992) in 124 sediment samples
high value may reflect the influence of the nearby sulfide ore
from the southern Beaufort Sea continental shelf, and 4.17
body (the Nanisivik mine started operation in October 1976).
±0.72 in 26 sediment samples from Hudson Bay. In sedi-
No dependence of Hg concentration on source rock geology
ments from the Beaufort Sea shelf, concentrations ranged
was detected from Greenlandic sediments (Loring and As-
from 9.1 to 20.1 mg/kg, corresponding to 0% and 100%
mund 1996), where the average concentration was 0.045
mud, respectively. The highest concentration of Pb in Arctic
±0.045 mg/kg.
sediments far from point sources is 53.0±2.2 mg/kg in the
Hg concentrations generally correlate negatively with se-
surface layer of three sediment cores from Ålesund, Nor-
diment grain size. On the Beaufort Sea shelf, Hg ranges from
way. This is probably of anthropogenic origin as the same
0.01 to 0.1 mg/kg in sediments containing 0% to 100% mud
cores had a concentration of only 22 mg/kg at a depth of
(Macdonald and Thomas 1991). Similarly, Loring (1984) re-
40-45 cm. The highest reported concentration in sediments
ported a grain size-Hg dependence for sediments from the
believed to be natural is 43.5±2.2 in sediments from south-
Canadian Arctic in the concentration range 0.04±0.014 to
Chapter 7 · Heavy Metals
421
Hg, mg/kg dw
7.6.5.3. Microorganisms
0
0.02
0.04
0.06
0.08
0.1
0.12
0.14
0.16
0
No data were found on the concentrations of heavy metals
1
in microorganisms from the Arctic.
2
3
4
7.6.5.4. Algae
5
6
Few data have been published on heavy metals in Arctic
7
algae. Data, primarily for Cd and to a lesser extent Pb, Hg,
8
9
and Se, are available for five intertidal species in Greenland,
10
but the geographical extent is very limited (Annex Table
11
7·A11). Additional data in algae are available only on Pb for
12
13
Canada and Norway.
14
15
Lead
16
17
The Greenland data do not indicate differences in Pb levels
18
between the three species analyzed for Pb. Lead concentra-
19
tions vary by a factor of about two, with geometric mean val-
20
21
ues in the growing tips ranging from 0.175 to 0.360 g/g dw.
22
No geographical differences within Greenland are indicated.
North Pole
23
In algae from Canada and Norway (Svalbard), Pb concentra-
24
25
Ålesund
tions are 3-5 times higher than those from Greenland. It is
26
suspected that these differences do not reflect real geographi-
27
Eastern
cal differences of baseline levels, but rather differences in sam-
28
Hudson Bay
29
ple preparation, analytical techniques, or species composition.
Central West
30
Lead concentrations are higher in whole algae plants than
Greenland
31
in growing tips (Annex Table 7·A11). No significant within-
32
33
year variations have been found in areas not affected by lo-
34
cal pollution in Greenland (Riget et al. 1995). Brown algae
35
Eastern
have been used as bioindicators of heavy metal pollution
36
Hudson Bay North Pole
37
from mining operations in Greenland, and the algae have
37.5
shown to be very good indicators of local geographic differ-
Depth
Central West
Greenland
Ålesund
ences and temporal trends of Pb pollution (Figure 7·38, Jo-
cm
hansen et al. 1991). At the Black Angel mine in northwest
Greenland, Pb in seaweed was elevated in a wide area (up to
Figure 7·37. Concentrations of Hg in Arctic marine sediment cores.
30 km from the mine), with values up to 50 g/g near the
(Source of data: Gobeil and Macdonald unpubl., North Pole; Norwegian
Institute for Water Research, Ålesund; Lockhart unpubl., eastern Hudson
Pb, µg/g dw
Bay; Dietz et al. 1997b, central West Greenland).
50
0.07±0.039 mg/kg. There are several data sets that indicate
widespread accumulation of Hg in surficial Arctic sediments
40
(Lockhart unpubl., Dietz et al. 1997b, OSPARCOM/NIVA
unpubl. data, Gobeil and Macdonald unpubl.; Figure 7·37).
30
The enrichment of Hg occurs in the upper 2 to 10 cm of the
sediments, even at the North Pole. This phenomenon could
indicate global scale input of Hg to the marine environment
20
in recent times. More comprehensive investigations of Arctic
sediments are required, however, before definite conclusions
10
can be drawn about the nature and source of the observed
enrichment.
0
Arsenic
1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995
Arsenic is not a part of the AMAP sediment monitoring
Figure 7·38. Temporal trend of Pb concentrations in growing tips of sea-
program, but data on As has often been reported for Arctic
weed (Fucus vesiculosus) at the Black Angel lead-zinc mine in West Green-
land. The background value is estimated to be 0.2 µg/g dw. (After Johan-
sediments. Loring et al. (1995) and Maage et al. (1996) re-
sen et al. 1997).
port values for As in sediments of the Pechora Sea and the
Barents Sea. The concentrations ranged from 2.0 to 308
mine (Johansen et al. 1997). Local elevations in Pb concen-
mg/kg. Loring et al. (1995) found high As concentrations
tration were also observed at the cryolite mine in south
in sediments contaminated with radionuclides from local
Greenland (Johansen et al. 1995), at the Pb-Zn mine in East
sources (e.g., near the nuclear weapons test site at Novaya
Greenland (Johansen et al. 1985), and at the Nanisivik Pb-
Zemlya). However, subsequent work by both Maage et al.
Zn mine at Strathcona Sound in the Canadian Arctic (As-
(1996) and Loring (pers. comm.) shows that As enrichment
mund et al. 1991).
is a common feature in sediments of the Barents and Pe-
chora Seas. The explanation for this enrichment is probably
Cadmium
natural and geological, however some attention might be
The concentration of Cd found in algae from Canada range
given to the potential effects of As accumulation in higher
from 0.7 to 1.8 g/g dw. The more comprehensive dataset
avians and mammals.
from Greenland shows a wider range, from 0.142 to 3.503
422
AMAP Assessment Report
% of
maximum level
100
90
80
70
60
50
40
30
20
10
0
Bladder wrack
Knotted wrack
Fucus disticus
Deep sea prawn
Blue mussel
Sculpin liver
Inner fjord
Central fjord
Outer fjord
Figure 7·39. Local differences in Cd concentrations in sedentary and stationary algae and biota from Nuuk and Uummannaq fjords, central West Green-
land (after Dietz 1995). Levels are plotted as the percentage of the maximum concentrations.
g/g dw (Annex Table 7·A11). Concentrations in knotted
tion in blue mussels increases slightly with increasing shell
wrack (Ascophyllum nodusum) are from three to six times
length (Riget et al. 1996). Older data (1983-1990) from
lower than in bladder wrack (Fucus vesicolosus) and the
southwest Greenland show Pb concentrations from 0.142 to
other wrack species, Fucus distichus, but are similar to the
0.476 g/g. Recent AMAP data from Greenland are similar
concentration reported for an unidentified kelp species from
to the lowest value of the older Greenland data, ranging
Svalbard.
from 0.072 to 0.155 g/g at Qeqertarssuaq (central West
No geographical differences on a regional scale are appar-
Greenland) and from 0.124 to 0.188 g/g at Nanortalik
ent. However, important local geographical differences in Cd
(south Greenland). Norwegian samples of blue mussels col-
concentrations were detected in Greenland. The highest con-
lected from 1984 to 1993 are in the range of 0.101-0.483
centrations were found near the open sea and the lowest con-
g/g, which is similar to the range observed in the same per-
centrations in the inner reaches of large fjords, with levels dif-
iod in Greenland. Old Icelandic data (1978) are relatively
fering by up to a factor of 5 (Riget et al. 1993) (Figure 7·39).
high (0.56-0.805 g/g), whereas Icelandic data from 1990 to
This agrees with Russian observations showing that the dis-
1992 are lower (range 0.019 to < 0.195 g/g). Older Green-
solved Cd concentration increases with the salinity in the Ob
land data (1984-1985) for other bivalve species are in the
and Yenisey estuaries (Dai and Martin 1995). Cadmium con-
same range (0.069-0.186 g/g) as the 1994 AMAP data for
centrations have also been found to vary seasonally. In Green-
blue mussels. Data on bivalves from areas outside Green-
land, Cd concentrations in bladder wrack in February were
land, except Iceland, cannot be directly compared with the
about a factor of 3 higher than in August (Riget et al. 1995).
Greenland data, as they appear to correspond to a different
At the three former mine sites in Greenland, Cd levels in
tissue. Concentrations in Canada and Russia seem to be
seaweed were not elevated above background levels, although
higher than in Greenland, but these data may not be reliable
Cd was higher than background in ore and mine waste (Jo-
and require confirmation and intercalibration.
hansen et al. 1985, 1991, Hansen and Asmund 1986).
At the Black Angel mine in northwest Greenland, Pb con-
centrations in blue mussels were elevated over background
Mercury and selenium
levels in a wide area (up to 30 km from the mine), with val-
Few data are available on Hg and Se concentrations in algae
ues exceeding 1000 g/g nearest the mine (Figure 7·40; Jo-
(Annex Table 7·A11). Selenium levels seem not to differ
hansen et al. 1991). Local elevations were also observed in
much (concentrations range from < 0.2 to 0.6 g/g dw),
blue mussels at the cryolite mine in south Greenland (Johan-
whereas larger differences have been reported for Hg (< 0.01
sen et al. 1985), in three bivalve species at the Pb-Zn mine in
g/g dw in kelp from Canada to more than 1 g/g dw in
East Greenland (Hansen and Asmund 1986), and in clams at
kelp from the Kara Sea, Russia). Data, however, are too few
and the uncertainty in the Kara Sea data is too high to reach
firm conclusions about possible geographical differences.
Black Angel
mine
250
81
7.6.5.5. Invertebrates
32.8
89
70 75
51.8
Data (Annex Table 7·A12) are available mainly for bivalves,
100
67
50
amphipods, and decapods. The following description of met-
25
al concentrations in invertebrates will concentrate on these
8
9.8
groups.
1.8
0
3
6 km
Pb, µg/g ww
Lead
In bivalves, most data are available for the blue mussel
Figure 7·40. Lead levels in blue mussels (Mytilus edulis) in the fjord out-
side the Black Angel lead-zinc mine in West Greenland in 1986-87. Values
which has been selected as an indicator species in the AMAP
on the fjord coast are measured concentrations. Estimated concentrations
program. Data from Greenland show that the Pb concentra-
in the fjord are indicated by isolines. (After Johansen et al. 1991).
Chapter 7 · Heavy Metals
423
the Nanisivik Pb-Zn mine at Strathcona Sound in the Can-
Cd concentrations in crustacea increase with the size (age)
adian Arctic (Asmund et al. 1991). At the Black Angel mine,
of the animals. In the amphipod Parathemisto libellula, Cd
local elevations of Pb were also observed in the deep-sea
concentrations are approximately twice as high in large as in
prawn (Johansen et al. 1991).
small animals (Annex Table 7·A12).
Data on Pb concentrations in crustacea, primarily Green-
For decapods, data are available mainly for Greenland.
landic deep-sea prawn, indicate higher values in heads and
Five species have been analyzed. Analyzes were either car-
shells than in muscle by a factor of 2 to 10, and higher in
ried out on whole animals or on the muscle and remaining
small than in large prawns. Geographical differences were
parts (heads and shells) separately. Cadmium levels are high-
not apparent, and Pb levels were generally low, with geo-
est in heads and shells (referred to as shell), intermediate in
metric mean values ranging from 0.008 to 0.050 g/g in
whole animals, and lowest in muscle. Shell values range be-
muscle, and 0.070-0.181 g/g in heads and shells. Compar-
tween 0.344 and 3.93 g/g ww, with lowest levels in small
able Pb concentrations were found in another shrimp species
deep-sea prawns from Nuuk and highest in large individuals
(Eualus belcher) from Greenland.
of the same species from Uummannaq. Whole animal values
Data on Pb in crustacea from other Arctic countries are
range between 1.28 and 7.91 g/g, with lowest levels in
generally very limited. The few available Russian data were
deep-sea prawns from Maniitsoq and highest in large Bikini
higher than other values, except for recent Norwegian data
prawns from Avanersuaq.
on meat from the deep-sea prawn from Canada and Nor-
Concentrations of Cd in muscle range between 0.013 and
way. Here levels are similar to what is found in Greenland.
0.046 g/g. The lowest concentrations are found in deep-sea
Lead concentrations in Arctic zooplankton from the Green-
prawns from Nuuk (both large and small) and small prawns
land Sea and Fram Strait appear high in some cases.
from Paamiut. The highest muscle levels are seen in large
For other invertebrate groups, data are limited but in-
deep-sea prawns from Uummannaq. The limited data from
clude a few results on polychaetes, echinoderms, and ascidi-
Canada and Norway indicate Cd levels similar to those in
ans from Russia. They cannot be compared with other areas
Greenland.
directly, but some of the Pb values appear to be relatively
The local geographical differences of Cd levels that have
high.
been found for seaweed and bivalves are also observed in
crustaceans. In deep-sea prawns the highest Cd concentra-
Cadmium
tions occur in samples from the open sea and the lowest in
Most of the mollusk data obtained prior to 1994 are for blue
samples collected in the inner regions of large fjord systems
mussels from three areas in West Greenland and Norway.
in Greenland (Figure 7·39); the observed levels differed ap-
Concentrations range from 0.458 g/g in small mussels from
proximately by a factor two (Johansen et al. 1991).
Uummannaq to 1.065 g/g in large mussels from Paamiut.
For the remaining groups of crustacea (copepods, ostra-
There is no indication of regional differences within Green-
cods, mysids, isopods, and euphausiids) few data are avail-
land (Annex Table 7·A12); however, local geographical dif-
able (Annex Table 7·A12). In general they also have high
ferences have been found. Similar to the situation for algae,
concentrations, but with large variations, from < 0.015 g/g
the highest Cd concentrations in blue mussels are seen in the
in euphausiids to 6.62 g/g ww in copepods.
open sea and the lowest in the inner regions of large fjords
For other invertebrate groups, annelids, echinoderms, as-
(Riget et al. 1996) (Figure 7·39). The highest Cd concentra-
cidians, and chaetognaths from Russia, Canada, and Nor-
tions are found in the largest mussels (Riget et al. 1996).
way, Cd data are sparse, and comparisons cannot be made,
The 1994 AMAP data from south Greenland (Nanor-
except to note that Cd levels in these groups seem to be low-
talik) range from 0.561 to 0.763 g/g and are at the same
er than in mollusks and crustacea.
level as the older Greenland data, whereas the new data
At the three former mine sites in Greenland, Cd concen-
from central West Greenland (Qeqertarssuaq) are somewhat
trations in invertebrates in most cases do not exceed back-
higher (0.692-1.254 g/g).
ground levels (Johansen et al. 1985, 1991, Hansen and As-
The data on blue mussels from Iceland indicate that Cd
mund 1986).
concentrations there are similar to or lower than those in
Greenland, whereas they are lower in blue mussels from
Mercury
Norway (0.124-0.353 g/g).
Most data are for bivalves from Greenland and Norway and
In two other bivalve species (a cockle, Serripes groen-
for decapods from Greenland.
landicus, and green crenella from northern and East Green-
In bivalves from Greenland, Hg levels are similar among
land), Cd concentrations are similar to those found in the
species and regions (Annex Table 7·A12). In soft tissue, the
blue mussel, whereas levels in Iceland scallop from north-
range is from 0.011 g/g in a cockle to 0.020 g/g in an Ice-
west Greenland are significantly higher, by a factor of 3-6.
land scallop. The 1994 AMAP data from Greenland fit well
Cadmium data in other bivalve species and in gastropods
into this range with levels from 0.014 to 0.017 g/g, and
from other countries are sparse, but the levels generally fall
similar values are found in blue mussels from Iceland. Con-
within the ranges observed in Greenland.
centrations of Hg in bivalve samples from Norway are simi-
Many data are available on Cd in crustacea, especially
lar (< 0.009-0.033 g/g) to those found in the Greenland
amphipods and decapods, but almost all are from Canada
samples.
and Greenland (Annex Table 7·A12). Some of the data
Although tissues and species are not similar, Hg values
were reported on a dry weight basis and some on a wet
for bivalves from Russian and Canada are similar to those
weight basis; a conversion factor of 0.25 for whole crustacea
found in Greenland and Norway. In general, Hg concentra-
(Dietz et al. 1996) has been used to convert to wet weight
tions in bivalves are low compared with those found in fish,
basis for comparisons.
seabirds, and marine mammals from the Arctic.
In general, high Cd concentrations are found in crustacea,
Although higher than in bivalves, Hg levels are low in
but with large variations, even within the same species. In
crustacea. In decapods from Greenland whole animal values
amphipods, Cd concentrations in whole animals range from
range between 0.023 and 0.258 g/g, with the lowest levels
0.11 to 4.60 g/g ww, with most values exceeding 1 g/g.
occurring in small Bikini prawns from Avanersuaq and the
There are no systematic geographical trends in the data.
highest in large deep-sea prawns from Baffin Bay.
424
AMAP Assessment Report
The Hg concentration is higher in whole large decapods
Pb, µg/g ww
than in small animals. No systematic geographical differ-
1.2
ences in Hg levels in decapods have been reported.
In copepods and ostracods, Hg concentrations seem to
1.0
be at the same level as in decapods, whereas in the remain-
ing crustacea groups (isopods, amphipods, and euphausiids)
and in annelids, echinoderms, and ascidians, data are few
0.8
but Hg levels appear lower than in the other crustacea
groups (Annex Table 7·A12).
0.6
Selenium
Most results are from Greenland, and almost all are for bi-
0.4
valves and decapods. In three bivalve species (Iceland scal-
lop, green crenella, a cockle), all from the same area in
0.2
northern Greenland, Se concentrations range from 0.326 to
0.869 g/g (Annex Table 7·A12). The level in the green
0
crenella is about twice as high as in the two other species.
1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995
The 1994 AMAP data on Se in blue mussels from two
regions in Greenland are at the same level as in green cre-
Figure 7·41. Temporal trend of Pb concentrations in liver tissue of spotted
wolffish (Anarhichas minor) at the Black Angel lead-zinc mine in West
nella, ranging from 0.53 to 0.94 g/g. The highest Se con-
Greenland. (After Johansen et al. 1996).
centrations are found in small blue mussels, decreasing by
one third in the largest mussels.
were higher in East Greenland than in West Greenland (Ri-
Selenium concentrations are higher in decapods than in
get et al. 1997a).
bivalves. As is the case for Cd, Se appears to be concentrated
in the heads and shells and not in the muscle of the animals.
Point sources. At the Black Angel mine in northwest Green-
The concentrations of Se in whole decapods from Greenland
land, local elevated Pb concentrations were measured in the
range between 1.09 and 3.91 g/g. The lowest concentra-
bone of shorthorn sculpin and spotted wolffish (Anarhichas
tions are found in small Bikini prawns from Avanersuaq and
minor). Johansen et al. (1985) report Pb concentrations of 5
the highest in large Sclerocrangons from Ittoqqortoormiit.
g/g in the bones of capelin (Mallotus villosus); concentra-
Although Se concentrations are higher in whole large crus-
tions were lower in liver and muscle tissue. A general decline
taceans than in small, differences are not large. No geo-
has been observed for a number of species due to improved
graphical differences on a regional scale are evident in the
treatment of mine waste (e.g., Figure 7·41). Elevated levels
existing data.
of Pb in fish were not observed at the cryolite mine in south
The few Se data in other animal groups (gastropods, am-
Greenland (Johansen et al. 1985), whereas Agger et al. (1991)
phipods, and ascidians) fall within the range of those found
report enrichment of Pb concentrations in the Pb-Zn mine in
in bivalves and decapods.
East Greenland.
Cadmium
7.6.5.6. Fish
Cadmium data are available for a substantial number of fish
The concentrations of Pb, Cd, Hg, and Se in marine fish
species in muscle, liver, kidney, bone, bile, spleen, and go-
species are presented in Annex Table 7·A13.
nads, but most are for muscle and liver tissue. The concen-
trations of Cd in fish muscle are generally below the detec-
Lead
tion limit (< 0.001-0.05 g/g). Cadmium levels in liver (gen-
Pb data are available for several fish species. Most data are
erally up to 1.0 g/g) were higher than in muscle. For wolf-
for muscle and liver tissue, a few are for bone, and one is for
fish, cusk (Brosme brosme), redfish (Sebastes marinus), short-
spleen and gonads. Generally, Pb concentrations in fish mus-
horn sculpin, and Greenland halibut (Rienhardtius hippo-
cle are very low (< 0.002-0.05 g/g). Where muscle and liver
glossoides), the liver values were even higher, between 1-12
have been analyzed, the levels in liver are higher than those
g/g. One of the reasons for these high Cd levels may be the
in muscle up to 0.2 g/g, mainly in fish species from Ork-
long life span of some of these fish species, especially wolf-
dalsfjorden/Trossavika, Norway which may be contami-
fish, Greenland halibut, and redfish. Stange et al. (1996) also
nated. The few data of Pb in bone tissue indicate levels simi-
found higher levels in redfish than in fish species with a short-
lar to those in liver. Many of the older data are higher than
er life span. Extremely high values were found in livers of
newer results; they are considered to be overestimates be-
Pacific herring (Clupea harengus) and broad whitefish
cause of analytical problems which were prevalent when the
(Coregonus nasus) from Tuktoyaktuk Harbour (30.6 and
samples were analyzed (Dietz et al. 1996, 1997b).
40.3 g/g). The levels of Cd in kidney were lower than in
liver for most of 13 Greenland species, below or just above a
Age accumulation. The data available for fish are too few to
value of 0.1 g/g. The highest value was found in the kidney
allow analysis for age/size dependence. Dietz et al. (1997a),
of a cusk (4.2 g/g).
Riget et al. (1997b), and Bohn and Fallis (1978) found no
Cadmium levels in fish bone from Greenland were all
correlation between fish size (age) and Pb concentrations.
< 0.015 g/g. The bile and spleen of a small number of
Greenlandic fish species contained mean Cd concentrations
Geographical differences. The limited data available for
from < 0.001 to 0.4 g/g. The highest individual values were
Pb in marine fish do not allow for general conclusions
found in an eelpout (Lycodes eudipleurostichus) from Ittoq-
regarding possible geographical differences. However, re-
qortoormiit (0.643 g/g and 2.06 g/g in spleen and bile,
cent investigation carried out under the AMAP program
respectively). Gonads from six fish species were analyzed,
showed that Pb levels in livers of Arctic cod (Boreogadus
and ranged from < 0.001 to 0.33 g/g in Arctic cod from
saida) and shorthorn sculpin (Myoxycephalus scorpius)
Pangnirtung.
Chapter 7 · Heavy Metals
425
Age accumulation. There is no solid evidence that the con-
centration of Cd increases with size of fish. However, Dietz
et al. (1997a) and Riget et al. (1997b) found a tendency of
increasing Cd concentrations with size in a limited portion
of the Greenlandic data, as did Bohn and Falls (1978) for
shorthorn sculpin. On the other hand, Hellou et al. (1992b)
found Cd concentrations in Atlantic cod (Gadus morhua)
were negatively correlated with size.
Kong Oscars
Avanersuaq Fjord
0.91 1.73
Grise Fjord*
Geographical differences. Cd levels in liver of Atlantic cod,
0.445
Pangnirtung
Greenland cod (Gadus ogac), redfish, shorthorn sculpin,
0.445
and especially Arctic cod can be compared over broad geo-
0.83
graphical areas. No clear geographical differences are evi-
Greenland Sea
0.39
dent as the variability of the data on a local scale is often
Ittoqqortoormiit
Cumberland
0.39
0.15
similar to that on a regional scale. However, the data indi-
0.66
Sound
0.37
cate that Arctic cod from the Barents Sea have lower con-
Barents Sea
0.21
centrations of Cd than Arctic cod from other areas, Figure
Jan
Mayen
7·42. Samples from Arctic cod and shorthorn sculpin col-
Nanortalik Denmark
Strait
lected in 1994 and 1995 showed higher Cd contents in liv-
Cd in liver
ers of individuals from northwest Greenland compared
µg/g ww
with other Greenland areas (Riget et al. 1997a). Local dif-
1.0
0.8
ferences have been found in the Cd content of livers of
0.6
shorthorn sculpin from some Greenland fjords: samples
*
0.4
Converted from dw til ww assuming a dry weight of
collected in the open sea can contain Cd concentrations up
30%, and converted from geometric mean to arithmetic
0.2
mean assuming a correction factor of 1.30.
to three times higher than those in sculpin collected from
0
the inner regions of large fjords (Figure 7·39) (Asmund et
Figure 7·42. Circumpolar distribution of Cd levels in liver tissue of Arctic
cod (Boreogadus saida). Arithmetic mean concentrations. (Sources of data:
al. 1988).
see Annex Table 7·A13).
Point sources. At three mine sites in Greenland (Black An-
gel Pb-Zn mine, cryolite (Al) mine in south Greenland, and
Pb-Zn mine in East Greenland), monitoring data suggest no
elevation of Cd in fish above background (Johansen et al.
1991, Hansen and Asmund 1986, Agger and Johansen 1992,
Barrow
Johansen et al. 1995, Asmund et al. 1988).
Strait
0.04
Resolute Bay
Mercury
0.02
0.04
Cambridge Bay
Data are available for Hg in muscle, liver, kidney, bile, spleen,
Avanersuaq
0.41
and gonads of fish; most data are for muscle and liver. In
Arctic Bay
muscle the concentrations of Hg are generally in the range
0.02
Pangnirtung
of 0.01-0.1 g/g. In some species, such as Greenland hal-
0.03
ibut, two eelpout species, Arctic cod, and shorthorn sculpin,
0.03
Greenland Sea
0.01
Hg concentrations are higher, between 0.1 and 0.2 g/g.
Ittoqqortoormiit
Cumberland
0.01
The highest values in fish muscle were found in Arctic cod
0.02
0.02
Sound
0.01
from Kong Oscars Fjord in East Greenland (0.882 g/g).
Jan
0.19
Barents
Mayen
Sea
Concentrations in liver were lower than in muscle, between
0.12
Kong
Oscars Fjord
0.01 and 0.06 g/g for most species. The highest value of
Nanortalik Denmark
Hg in liver tissue was found in samples of Arctic cod from
Hg
Strait
µg/g ww
Upernavik (0.192 g/g).
0.05
In kidneys, Hg concentrations ranged between 0.013 and
0.04
0.078 g/g, except in Greenland cod from Cambridge Bay
0.03
0.02
(0.17 g/g). The concentration of Hg in the spleen and bile
0.01
of Greenlandic fish ranged from 0.011 to 0.122 g/g and
0
from < 0.005 to 0.067 g/g, respectively. In gonads of nine
Figure 7·43. Circumpolar distribution of Hg levels in muscle tissue of Arc-
fish species Hg concentrations ranged from 0.01 to 0.09 g/g.
tic cod (Boreogadus saida). Arithmetic mean concentrations. (Sources of
data: see Annex Table 7·A13).
Age accumulation. Dietz et al. (1997a) and Riget et al.
Hg in fish from the Barents Sea and Greenland Sea are lower
(1997b) found a clear, positive correlation between Hg
than in the other areas; however, the sampling periods dif-
concentrations in muscle, liver, and kidney and size of fish.
fered considerably among areas, which may have influenced
Stange et al. (1996) report a similar relationship for fish
the levels found (Figure 7·43). Some of the highest values
from the North Atlantic.
were found in fish from the northern areas of eastern Can-
ada as well as western and eastern Greenland.
Geographical differences. There are very few fish species
which have been sampled on a circumpolar scale and ana-
Temporal trends. Very limited information is available on
lyzed for Hg. An exception is Arctic cod for which the con-
temporal trends in marine fish. A study from two locations
centration of Hg in muscle tissue is available for the eastern
in the Baltic (Landsort and Utlängen) south of the Arctic
Canadian Arctic, West and East Greenland, Jan Mayen, the
area (58°45'N), however, have shown an increasing trend
Greenland Sea, and the Barents Sea. The concentrations of
over the period from 1980 to 1993 (Bignert et al. 1995).
426
AMAP Assessment Report
Selenium
Geographical differences. No seabird species has been ana-
Data on the Se content of 20 fish species are available. In
lyzed for Pb on a circumpolar basis. In general, however,
muscle, concentrations range from below detection (0.2 g/g
seabirds from Russia have a higher Pb level than seabirds
in the Greenland data) to 1.01 g/g in shorthorn sculpin
from other areas. This situation may reflect differences in
from East Greenland. In liver, concentrations are clearly
sampling and analytical methodology rather than real differ-
higher than in muscle; the highest values are reported for At-
ences in concentrations. The 1994 AMAP results on gulls
lantic wolffish, Greenland halibut, and golden redfish, all
from four regions in Greenland indicate higher Pb levels in
from Nuuk, West Greenland (~3 g/g), for redfish from
East Greenland than in other regions of Greenland (Riget et
Denmark Strait (4.87 g/g), and for Pacific herring from
al. 1997a).
Tuktoyaktuk Harbour (3.26 g/g). The concentrations of
Se in kidneys are similar to or marginally higher than those
Point sources. At the Black Angel mine in northwest Green-
found in liver. Long rough dab from south Greenland had
land, Pb levels in muscle, liver, and kidney from seabirds
the highest concentration, in kidney tissue, 6.30 g/g. The
(common eider; glaucous gull, Larus hyperboreus; Iceland
Se concentration in spleen of nine Greenlandic fish species
gull; and black guillemot, Cepphus grylle) were not ele-
ranged from 0.39 g/g in cusk from south Greenland to
vated, whereas Pb concentrations in bone were two to ten
5.25 g/g in Arctic cod from Ittoqqortoormiit, East Green-
times higher than at reference sites (Agger and Johansen
land. In four fish species, all from Nuuk, West Greenland,
1992).
the levels of Se in bile range from below detection (< 0.2
g/g) to 0.74 g/g. Levels of Se in gonads were generally
Cadmium
low except for Arctic sculpin from Tuktoyaktuk Harbour.
Most data for Cd in seabirds are from Greenland, Canada,
and Norway. Some data are available for Iceland and Russia
Age accumulation. Concentrations of Se in liver tended to
(Annex Table 7·A14). In muscle tissue, some extreme high
decrease with increasing length of fish. Stange et al. (1996)
values (up to 28 g/g) are found in fulmar from Svalbard,
report a similar trend for Atlantic cod but not for redfish.
Norway. High levels in muscle tissue are also found in sea-
No correlation between the concentration of Se in muscle or
birds from Lancaster Sound (northern fulmar, 2.59 g/g;
kidney tissue and fish length is apparent in the data (Dietz et
kittiwake, Rissa tridactyla, 1.62 g/g; Brünnichs guillemot,
al. 1997a, Riget et al. 1997b).
Uria lomvia, 1.55 g/g). Most other values of Cd in muscle
are below 0.6 g/g.
Geographical differences. Data for the Se concentration in
Cadmium concentrations generally decrease in the order
liver of Arctic cod exist for the eastern Canadian Arctic,
of kidney > liver > muscle. In liver, Cd concentration ranges
West and East Greenland, Jan Mayen, the Greenland Sea,
from below 1 g/g in yearlings of most seabird species to
and the Barents Sea, but no clear geographical differences
about 10 g/g in older birds, except for several seabird spe-
are apparent.
cies from Svalbard where values up to 30 g/g in fulmar
have been reported. High values (approx. 20 g/g) were also
found in northern fulmar and Brünnichs guillemot from Lan-
7.6.5.7. Seabirds
caster Sound, Canada and in glaucous gulls from Avaner-
Lead
suaq, northwest Greenland. In kidney tissue, yearling seabirds
Several seabird species were analyzed for Pb, mostly from
have low levels, ~0.1 g/g. The highest values are found in
Greenland, Canada, and Norway (Annex Table 7·A14).
older birds with levels up to 80 g/g in glaucous gull from
Few data are available from Iceland and Russia. In seabird
the northern part of West Greenland. In five seabird species,
muscle, some extreme high values (up to 26 g/g) are found
Cd levels in bone tissue are low, ranging from below detec-
in several species from Svalbard, Norway. Whether these
tion to 0.112 g/g in black guillemot from Uummannaq,
high values are real is unknown. It is necessary to evaluate
West Greenland.
the data for Pb in the breast tissue of birds with caution,
particularly if they were killed by lead shot. All other Pb val-
Age accumulation. In cases where Cd concentrations are
ues are below 0.4 g/g, except for one specimen of king
given for different age groups (mostly from Greenland), the
eider (Somateria spectabilis) from Holman Island, Canada
levels generally increase with age (Nielsen and Dietz 1989).
which contained 1 g/g of Pb. The levels in liver and kidney
In the 1994 AMAP results from Greenland, Riget et al.
ranged from < 0.009 to 0.8 g/g. The highest concentrations
(1997a) found that Cd concentrations in liver of glaucous
in kidney are found in seabirds from Svalbard. The analyzes
gull and Iceland gull clearly increased with the age of the
of bone from five species of seabirds from Lancaster Sound,
gulls. However, Norheim (1987) found no differences be-
Canada and for the common eider (Somateria molissima)
tween juvenile and adult birds from Svalbard, including
from Uummannaq, West Greenland suggest that the Pb con-
glaucous gull, Brünnichs guillemot, little auk (Alle alle),
centrations in Lancaster Sound are abnormally high, possi-
and common eider.
bly because these data are old (1977).
Geographical differences. No clear geographical differences
Age accumulation. In general, few data on the Pb concentra-
are evident. However, in all cases the concentrations of Cd
tions in seabirds are suitable for evaluating the relationship
in birds from Lancaster Sound are higher than that in birds
between age and accumulation. Data for 25 gulls sampled in
from Greenland. The levels in the birds from Svalbard are
1994 in four areas of Greenland as part of the AMAP pro-
for some species lower or similar to those from Greenland
gram, however, were found suitable (Riget et al. 1997a) for
and for other species higher than those from Greenland.
addressing this question. The Pb levels in livers of these gulls
Several comparisons of seabirds of the same age can be
were not correlated with the age of the birds. Limited data for
made for the Greenland area. The highest levels were found
composite samples of common eider from Pangnirtung, Sani-
in northwest Greenland (Nielsen and Dietz 1989, Dietz et al.
kiluaq, and Uummannaq, long-tailed duck and red-breasted
1996). In contrast, the levels in gulls sampled as part of the
merganser (Mergus serrator) from Inukjuak, and Iceland gull
1994 AMAP program were very similar from north to south
(Larus glaucoides) from Uummannaq support this conclusion.
in West Greenland (Riget et al. 1997a).
Chapter 7 · Heavy Metals
427
Point sources. At three mine sites in Greenland, results of a
Age accumulation. Very few data exist for assessing possible
monitoring program indicate no elevation of Cd in seabirds
age accumulation of Se in seabirds. Riget et al. (1997a),
relative to background levels (Hansen and Asmund 1986,
using new AMAP data, found an increase in Se with age in
Asmund et al. 1988, Johansen et al. 1991, 1995, Agger and
glaucous and Iceland gulls.
Johansen 1992).
Geographical differences. Few data are available for com-
Mercury
paring Se levels in seabirds from around the Arctic. Selenium
Hg concentrations generally decrease in the order liver >
and Hg concentrations are highly correlated and, as is the
kidney > muscle (Annex Table 7·A14). The highest concen-
case for Hg, the levels of Se in seabirds from Canada, Green-
tration in muscle tissue (2.19 g/g) was found in great cor-
land, and Svalbard are similar. Within Greenland, Nielsen
morant from Kangaatsiaq, West Greenland. This value is
and Dietz (1989) and Riget et al. (1997a) found a tendency
based on one specimen, which also had the highest concen-
toward higher concentrations in birds from northwest Green-
tration in liver and kidney. Other values range from below
land and northeast Greenland than in those from south
detection to about 1 g/g in red-breasted merganser from
Greenland.
Kangiqsualujjuaq, Canada, and glaucous gull from Ava-
nersuaq, northwest Greenland. Several seabird species have
7.6.5.8. Marine mammals
been analyzed for Hg in liver tissue. Levels in liver range
from 0.1 to 3 g/g. Except for a high value of 28.1 g/g in
A substantial amount of data is available on the concentra-
kidney tissue of one specimen of great cormorant from West
tions of metals in Arctic marine mammals. Because Cd and
Greenland, the levels were intermediate between the values
Hg bioaccumulate strongly and because Hg biomagnifies,
in liver and muscle, ranging from 0.053 g/g in one-year-old
much more emphasis has been placed on obtaining body
kittiwake to 2.33 g/g in older glaucous gulls. Seven seabird
burdens of metals in biota of higher trophic levels.
species have been analyzed for Hg in feathers. The highest
levels were found in kittiwake (5.5 g/g fresh weight).
Lead
Data on tissue concentrations of Pb in marine mammals
Age accumulation. Few data are available for assessing age
have been produced for ten species including polar bear.
accumulation of Hg in seabirds. Levels in older kittiwake,
Most data are from Canada and Greenland, some are from
Brünnichs guillemot, and black guillemot from Lancaster
Alaska and Norway, and almost none are from Iceland and
Sound, Canada and black guillemot from West Greenland
Russia. Concentrations of Pb range from below detection
were generally higher than those in younger birds. No accu-
(< 0.010 g/g) to 0.083 g/g in tissues of ringed seals. For
mulation with age was found in livers of glaucous and Ice-
other pinniped species (harp seal, Pagophilus groenlandicus;
land gulls from Greenland in the new AMAP data (Riget et
fur seal, Callorhinus ursinus; and walrus), they range be-
al. 1997a), or in five seabird species from Greenland (Niel-
tween 0.010-0.328 g/g. High mean values (up to 1.0 g/g)
sen and Dietz 1989).
are found in some polar bear tissue from Svalbard (Norheim
et al. 1992), but some of these results are old and question-
Geographical differences. No clear conclusion can be drawn
able. A recent study by Wagemann et al. (1996) includes
about geographical differences of Hg in seabirds across the
data for ringed seals, beluga whales (Delphinapterus leucas),
Arctic. Mercury levels in seabirds from Arctic Canada are
and narwhals and shows no Pb levels higher than 0.083
similar to those from Greenland (Nielsen and Dietz 1989),
g/g. Dietz et al. (1996) reviewed previous Pb data from
Svalbard, and northern Norway. Thompson et al. (1992a)
Greenland and found that mean Pb concentrations for seals
found higher Hg concentration in feathers of birds from
did not exceed 0.058 g/g. Differences between tissues are
northwest Iceland than those from northern Norway. Within
not as apparent for Pb as for other heavy metals (Annex
Greenland, Nielsen and Dietz (1989) and Dietz et al. (1996)
Table 7·A15).
found higher concentrations in birds from northwest and
northeast Greenland than in those from south Greenland.
Age accumulation. Wagemann et al. (1983, 1990, 1996)
Riget et al. (1997a) presented new AMAP data, and found
found no correlation between Pb levels and the size or age of
higher concentrations in livers of glaucous gulls from East
narwhals and belugas from Canada. A similar lack of corre-
Greenland but no elevated levels in northwest Greenland
lation has been reported for white-beaked dolphins (Lageno-
compared with south Greenland.
rhynchus albirostris) and pilot whales (Globicephala melae-
na) from Newfoundland by Muir et al. (1988). The data by
Temporal trend. Appelquist et al. (1985) used feathers to
Norheim et al. (1992) for Pb in liver and kidney of adult
test for a historical trend of Hg concentration in black guil-
and juvenile polar bears from Svalbard also show no corre-
lemot and Brünnichs guillemot from Greenland, and found
lation between age and tissue Pb concentrations.
the levels to be almost constant during this century, though
slightly increasing in black guillemot.
Geographical differences. Possible geographical differences
are difficult to evaluate as analytical problems may well ob-
Selenium
scure true differences because Pb levels in marine mammals
Concentrations of Se generally decrease in the order kidney
are low, and frequently below or close to the detection limit.
> liver > muscle. Several seabird species have been analyzed
Wagemann et al. (1996), however, found significantly higher
for Se in muscle, liver, and kidney tissues (see Annex Table
Pb concentrations in muscle and liver tissues of ringed seals
7·A14). The levels in muscle range from below detection to
from the western Canadian Arctic compared with those from
4.95 g/g in older kittiwake. In seabird liver, the levels
the eastern Canadian Arctic, but no differences in muscle
ranged from 1.03 in little auk to 21.4 g/g in common eider,
and liver of belugas from the northern Canada (Wagemann
both from Uummannaq, West Greenland.
et al. 1990, 1996). The levels of Pb were higher in belugas
Selenium levels in kidney ranged from 1.24 g/g in ivory
from Hudson Bay than those from the Mackenzie Delta.
gull from Uummannaq, West Greenland to 27 g/g in glau-
Also, belugas from St. Lawrence River, Quebec, were signifi-
cous gull from Ittoqqortormiit, East Greenland.
cantly higher (0.104-0.159 g/g) than belugas from the Arc-
428
AMAP Assessment Report
tic as a whole, which was attributed to anthropogenic input
Cd
of Pb. Lead levels in liver of dolphins off Newfoundland
µg/g ww
(Muir et al. 1988) are of the same order of magnitude, but
3.5
are lower in muscle. In liver of ringed seals from the north-
ern part of central West Greenland, Pb levels were
half those in seals from central East Greenland (Dietz et
3.0
al. 1996).
2.5
Temporal trends. Few reliable data are available to evaluate
East Greenland
Cd = 0.80 + 0.07 Age, r2 = 0.152, N = 38
temporal trends of Pb in marine mammal tissues. Wage-
mann et al. (1996) found no temporal trends in muscle and
2.0
Northeast Canada
liver of narwhal from the eastern Canadian Arctic over a
Cd = 0.65 + 0.05 Age, N = 52
span of 14 years (1979-1993).
1.5
Cadmium
Central northern Canada
Cd = 0.45 + 0.03 Age, N = 36
Data for Cd in marine mammal tissues are available for six
1.0
seal species, six whale species, and polar bears (Annex Table
Svalbard
juveniles, Cd = 0.3; adults, Cd = 0.6
7·A15). The concentrations vary considerably among tissue
0.5
types, but generally decrease in the order kidney > liver >
Northwest Canada
muscle.
Cd = 0.09 + 0.02 Age, N = 30
In ringed seals, the highest Cd concentrations (geometric
0
means) are found in animals from northwest Greenland
0
2
4
6
8
10
12
14
16
18
20
Age
(111 g/g in kidneys, 36.7 g/g in livers). There were no
years
clear differences among seal species. In whales, the highest
Figure 7·44. Age accumulation of Cd in liver of polar bears from different
concentrations (63.5 g/g) were found in kidneys of nar-
Arctic regions. (After Dietz et al. 1995a).
whals from Pond Inlet (Wagemann et al. 1983, 1996). Minke
whales (Balaenoptera acutorostrata) had the lowest kidney
age of four years in harbour porpoises from Greenland
concentrations (4.24 g/g), whereas belugas, pilot whales,
waters. In animals older than four years, muscle and liver
harbour porpoises, and white-beaked dolphins had interme-
concentrations reached a constant level, whereas kidney
diate concentrations. Few data are available for bowhead
levels decreased.
whales (Balaena mysticetus). Two adult specimens (Becker
An increase of Cd concentrations in polar bear tissues
et al. 1995b) had considerable Cd in the liver (17.5-18.9
with age has been documented by several authors (Norstrom
g/g). Other tissues such as blubber and skin (mattak) are
et al. 1986, Braune et al. 1991, Dietz et al. 1995) (Figure
very low in Cd, and in most cases the levels are close to or
7·44), but no evidence of a decrease in old bears has been re-
below detection.
ported.
The concentrations of Cd in polar bears are remarkably
low compared with seals and whales, even though the polar
Geographical differences. Concentrations of Cd in marine
bear is a top predator. This has been explained by the feed-
mammal tissues increase from west to east in the Canadian
ing habits of the polar bear. It feeds virtually exclusively
High Arctic for ringed seal, beluga, and polar bear (Nor-
on ringed seal blubber, which has a very low Cd concentra-
strom et al. 1986, Braune et al. 1991, Wagemann et al.
tion (e.g., Norstrom et al. 1986, Braune et al. 1991, Dietz et
1996). This trend could be extended to cover West Green-
al. 1995). Cadmium concentrations in kidneys, liver, and
land, for ringed seals and polar bears, but not for belugas
muscle of polar bears are 19.7, 1.67, and 0.024 g/g, re-
(Annex Table 7·A15 and Figures 7·45 to 7·47). Wagemann
spectively.
et al. (1996) explain the geographical trend within Canada
in terms of geological differences between the western and
Age accumulation. In Annex Table 7·A15, metal levels are
the eastern Canadian Arctic. The eastern part of Canada
compiled for a large range of ages. For geographical and
consists essentially of Precambrian igneous and metamor-
temporal comparisons, it is of major importance to assess
phic rocks with a higher Cd concentration than is found in
the relationship between age and metal levels. Cadmium is
the Postcambrian unmetamorphosed sedimentary rocks
virtually absent from the body at birth, both in humans and
characteristic of the western Canadian Arctic. The correla-
marine mammals (WHO 1992a, 1992b, Dietz et al. 1995,
tion between regional geology and levels in marine biota in-
1997a). The increase of Cd concentration with age in seals
dicates that natural sources may be the primary cause of the
has been documented by several authors (Sergeant and Arm-
high levels in the eastern Arctic. In Greenland, no significant
strong 1973, 1980, Wagemann et al. 1996). Dietz et al. (in
differences in the Cd levels of bottom sediment were found
press) also found a highly significant correlation of Cd with
for the different geological structures (Loring and Asmund
age, but the concentration increased only until the seal reached
1996). On the other hand, Cd levels are generally highest in
adulthood and decreased thereafter. Consequently, Cd con-
ringed seals and polar bears from northwest Greenland com-
centrations in old seals were one third lower than in young
pared with areas farther south (Dietz et al. 1996).
adult seals.
A five-fold difference in Cd tissue concentrations occurs
Both Wagemann et al. (1983, 1990, 1996) and Muir
across the Arctic and depends on the areas, species, and tis-
et al. (1988) have noted the increase in Cd concentrations
sues examined. Cadmium levels in ringed seals and polar
with age in whales from Canadian waters. In a study on
bears in central East Greenland are somewhat lower than in
minke whales, belugas, and narwhals from Greenland,
northwest Greenland (Avanersuaq); they are even lower
Hansen et al. (1990) showed that in eight out of 18 ani-
around Svalbard. On the east cost of Greenland (Ittoqqortor-
mals, Cd concentrations increased significantly with age
miit, Danmarkshavn, and Kong Oscars Fjord), only minor
and that middle-aged whales had the highest values. Palu-
differences in the concentrations of Cd found in ringed seal
dan-Müller et al. (1993) found that Cd increased until the
occur between areas, with no discernible north-south trend.
Chapter 7 · Heavy Metals
429
Shingle Point
40
Paulatuk
30
40
20
30
10
20
0
10
0
Sachs Harbour
40
Holman
30
20
20
10
10
0
0
Resolute
1987
1988
40
30
20
10
Eureka
Admiralty Inlet
0
40
40
30
30
20
20
10
10
0
0
Umiujaq
40
Inukjuak
40
Nanisivik mine
30
Avanersuaq
40
40
30
20
Salluit
30
30
20
40
10
20
20
10
30
0
10
10
0
20
0
0
10
1984
1994
0
Upernavik
40
Uummannaq
40
30
Danmarkshavn
Wakeham Bay
40
40
30
20
Kuujjuarpik
30
30
20
10
(Great Whale)
20
40
20
10
0
10
30
10
0
0
20
0
1989
1990
10
Kong Oscars Fjord
Kangiqsualujjuak
40
0
Qeqertarssuaq
(George River)
40
30
Svalbard
40
50
30
20
30
40
20
10
20
30
10
0
10
20
0
0
10
0
1986
1992
1993
Ittoqqortoormiit
40
30
Nanortalik
40
20
30
10
20
0
1986
1994
10
Cd, µg/g ww
0
Age :
0 y
0-1 y
2-4 y
5-10 y
10-15 y
>15 y
Undetermined
Figure 7·45. Distribution of Cd levels in liver tissue of ringed seal (Phoca hispida) of different ages (years). Plots show selected data (geometric mean val-
ues) from Annex Table 7·A15. (Sources of data: see Annex Table 7·A15).
Cadmium concentrations in ringed seals from the Arc-
Müller et al. 1993). These differences may be partly ex-
tic are many times higher than those reported for ringed
plained by differences in available food items. Species such
seal from the Gulf of Finland and the Gulf of Bothnia
as parathemisto, other crustaceans, and Arctic cod may be
(Helle 1981, Perttilä 1986, Frank et al. 1992; data not
important Cd sources in the Arctic. The higher levels in
included in Annex Table 7·A15). Specifically, concentra-
Arctic marine mammals may also be a consequence of
tions are approximately 15 times higher in muscle, 16-75
slower growth rates in the Arctic. Slow growing poikilo-
times higher in liver, and 24-42 times higher in kidney. Jo-
thermic organisms accumulate metals over a longer period
hansen et al. (1980) also concluded that Cd was high-
of time before being eaten (Muir et al. 1996). In Svalbard,
est in Arctic seals. Concentrations of Cd in harbour por-
conditions seem different. Although situated far north,
poises from Greenland waters also had ten times higher
Svalbard is strongly influenced by the relatively warm Gulf
Cd levels than did those from European waters (Paludan-
Stream, leading to faster growth of the lower food chain
430
AMAP Assessment Report
West Whitefish
Paulatuk
Tuktoyaktuk
Station
Shingle Point
20
20
20
20
15
15
15
15
10
10
10
10
Kendall Island
20
5
5
5
5
15
0
0
0
0
10
5
0
Hendrickson Island
20
15
10
5
0
1993
1994
Arviat
East Whitefish
20
(Eskimo Point)
20
15
15
10
Repulse Bay
Grise Fjord
10
5
20
20
5
0
15
15
1993
1994
0
10
10
East Whitefish Station
5
20
5
0
15
0
10
Sanikiluaq
(Belcher Island)
5
20
Coral Harbour
0
15
20
10
15
5
10
0
5
0
Upernavik
20
15
Umiujaq
10
Kuujjuarpik
20
5
(Great Whale)
15
Iqaluit
20
0
20
10
15
15
5
Kangaatsiaq
10
10
20
0
5
5
15
0
10
0
1993 1994
5
0
Cd, µg/g ww
Age :
0-1 y
2-4 y
5-15 y
15-25 y
>25 y
Undetermined
Figure 7·46. Distribution of Cd levels in liver tissue of beluga whale (Delphinapterus leucas) of different ages (years). Plots show selected data (geometric
mean values) from Annex Table 7·A15. (Sources of data: see Annex Table 7·A15).
organisms, which ultimately results in lower body burdens
approximately 14 years in the Canadian High Arctic. Hansen
of metals.
(1988) analyzed hair samples from seals and humans from
the 15th century and found the levels in humans were not
Temporal trends. Limited information is available on tempo-
significantly different from those today, whereas levels in seal
ral trends in marine mammals. Wagemann et al. (1996) is the
furs were significantly higher (2.6 times) now than in the old
only study addressing this question. They found no temporal
samples. However, the value of hair as a satisfactory indica-
trends of Cd in tissues of narwhal and beluga sampled over
tor of Cd exposure was questioned by the author.
Chapter 7 · Heavy Metals
431
n.d.
0.659
0.528
Chukchi Sea
Eastern
Beaufort Sea
n.d.
0.371
1.75
Amundsen
Spence
Western
Gulf
Bay
Beaufort Sea
0.877
0.213
0.780
1.09
1.40
0.18
S.W.
Hadley Melville
Bay Island
Northern
Cornwallis Island
Hudson Bay
1.23
0.6
Southern
Cumberland
Hudson Bay
Peninsula
Avanersuaq
0.952
1.18
1.21
Svalbard
Ittoqqortoormiit
Clyde River
Northern
Cd, µg/g ww
Baffin Island
1.0
Cape Mercy
0.8
0.6
0.4
0.2
0
Figure 7·47. Distribution of Cd levels (age adjusted to 6.9 years, Svalbard adults) in liver tissue of polar bear (Ursus maritimus). (Source of data: Braune
et al. 1991, Norheim et al. 1993, Dietz et al. 1995, 1996).
Point sources. Ringed seals caught near the Nanisivik Pb-Zn
Concentrations of Hg in seal and whale muscle frequently
Mine in Strathcona Sound and at a reference site in Admi-
exceed 0.50 g/g ww, particularly in older individuals be-
ralty Inlet had comparable Cd levels (Wagemann 1989).
cause of the accumulation of Hg with age. The highest mean
Similarly, at the three mine sites in Greenland noted earlier,
Hg concentrations (0.72 g/g in muscle, 32.6 g/g ww in
Cd has been monitored in ringed seals, and at no site was it
liver) are found in ringed seals in the western Canadian
elevated relative to reference sites (Hansen and Asmund
Arctic (Wagemann et al. 1996). Up to 219 g/g have been
1986, Asmund et al. 1988, Johansen et al. 1991, 1995, Ag-
measured in the liver tissue of a ringed seal from Sachs Har-
ger and Johansen 1992).
bour. This is also the population where the highest mean
values have been recorded (Wagemann et al. 1996; see Fig-
Mercury
ure 7·48, next page). Very high levels, 143 g/g on average,
Mercury data from eight seal species, eight whale species,
have been reported for bearded seal from the Amundsen
and polar bears are compiled in Annex Table 7·A15, show-
Gulf (Smith and Armstrong 1975, 1978). Mercury can be
ing, as for Cd, considerable differences among tissue types.
transferred from the mother to the fetus during gestation,
Concentrations generally decrease in the order liver > kidney
which is apparently not the case for Cd (Wagemann et al.
> muscle. In polar bears, however, the highest Hg levels are
1988). Almost all the Hg in muscle tissue is present as meth-
found in the kidney.
ylmercury, whereas in liver tissue methylmercury seldom ex-
432
AMAP Assessment Report
Single Point
50
40
Paulatuk
30
103
50
20
40
10
30
0
70
20
10
Sachs Harbour
50
0
40
30
Holman
20
50
40
10
0
30
1987
1988
20
10
Resolute
0
50
40
Eureka
30
50
20
40
10
30
Sanikiluaq
0
20
0
(Belcher Island)
10
50
Admiralty Inlet
Nanisivik Mine
50
50
40
Salluit
50
40
40
30
40
Avanersuaq
30
30
20
50
30
20
20
10
40
20
10
0
10
30
10
0
0
20
Inukjuak
0
Uummannaq
10
Svalbard
50
50
50
0
40
Wakeham Bay
1984
1994
40
Danmarkshavn
40
50
30
50
30
30
40
Upernavik
20
40
20
50
20
30
10
30
10
40
10
20
0
20
0
30
0
10
10
20
0
0
1989
1990
10
Qeqertarssuaq
0
50
Kong Oscars Fjord
50
Kangiqsualujjuak
40
40
(George River)
30
50
30
20
40
20
Umiujaq
10
50
30
10
Kuujjuarpik
0
40
20
(Great Whale)
0
50
Ittoqqortormiit
30
10
50
40
20
0
Nanortalik
40
30
10
50
30
20
0
40
20
10
30
10
0
20
0
10
Hg, µg/g ww
1986
1994
0
Age :
0-1 y
2-4 y
5-10 y
10-15 y
>15 y
Undetermined
Figure 7·48. Distribution of Hg levels in liver tissue of ringed seal (Phoca hispida) of different ages (years). Plots show selected data (geometric mean
values) from Annex Table 7·A15. (Sources of data: see Annex Table 7·A15).
ceeds 2 g/g, even when the total Hg concentration is high
spectively. In contrast, Hg levels in most toothed whales are
(Dietz et al. 1990, Figure 7·49). In the kidney tissue of adult
comparatively high. In livers of white-beaked dolphins, har-
seal and whale, organic Hg is 10-20% of total Hg, whereas
bour porpoises, narwhals, belugas, and pilot whales, maxi-
in polar bears it is 6% (Dietz et al. 1990, Figure 7·49). This
mum means are 0.831, 8.21, 10.8, 42.4, and 280 g/g, re-
fraction is consistently higher, up to 70%, in young seals
spectively. The extremely high values of total Hg in pilot
and walrus (Born et al. 1981, Wagemann et al. 1988, Dietz
whales were from samples taken in 1977 near the Faeroe Is-
et al. 1990).
lands, and are well above the concentrations measured sub-
In general, Hg levels are low in baleen whales. Maximum
sequently in that area (38 to 84 g/g). In those 1977 sam-
levels in livers of bowhead, fin (Balaenoptera physalus), and
ples, methylmercury levels up to 35±10 g/g were reported
minke whales do not exceed 0.30, 0.546, and 0.452 g/g re-
(Julshamm et al. 1987).
Chapter 7 · Heavy Metals
433
organic-Hg, µg/kg ww
Hg
10 000
µg/g ww
60
Northwest Canada
Hg = 8.8 + 5.95 Age, N = 23
50
1 000
Northern Alaska
40
young, Hg = 22.4; adults, Hg = 38.1
30
Northeast Canada
100
Birds
Hg = 4.4 + 1.76 Age, N = 81
Seals
Toothed whales
East Greenland
20
Hg = 5.3 + 0.71 Age, r2 = 0.334, N = 38
Baleen whales
Polar bears
Hudson Bay, Canada
10
10
Hg = 1.3 + 0.71 Age, N = 20
10
100
1 000
10 000
100 000
Western Alaska : young, Hg = 3.9; adults, Hg = 4.8
total-Hg, µg/kg ww
Svalbard : juveniles, Hg = 1.9; adults, Hg = 2.6
0
0
2
4
6
8
10
12
14
16
18
20
Age
years
organic-Hg, µg/kg ww
Figure 7·50. Age accumulation of Hg in liver of polar bears from different
10 000
Arctic regions. (After Dietz et al. 1995).
1978, Wagemann et al. 1996) and for whales in the Cana-
dian Arctic (Gaskin et al. 1972, 1979, Wagemann et al.
1983, 1990, 1996, Muir et al. 1988). Hansen et al. (1990)
1 000
found that Hg was positively correlated with age in muscle,
liver, and kidney of narwhals, in liver and kidney of belugas,
and in liver of minke whales. Paludan-Müller et al. (1993)
found that in harbour porpoises from Greenland waters, Hg
100
increased with age in muscle, skin, and perhaps also in kid-
Birds
Seals
ney until four years of age. By contrast, Hg appears to in-
Toothed whales
crease in whale liver throughout the entire lifetime of the an-
Baleen whales
imal. Julshamm et al. (1987) found that Hg concentration in
Polar bears
both the muscle and liver of pilot whales reached a plateau
10
once the whales reached a certain size. An increase of Hg
10
100
1 000
10 000
100 000
total-Hg, µg/kg ww
concentrations in the liver and kidneys of polar bears with
age has also been documented (Norstrom et al. 1986, Brau-
ne et al. 1991, Dietz et al. 1995; Figure 7·50).
organic-Hg, µg/kg ww
Geographical trends. Based on a sample size of two, the
10 000
Birds
Seals
concentration of Hg in ringed seals in Alaska appears to be
Toothed whales
high (1.52 and 3.52 g/g in a one- and a two-year-old ringed
Baleen whales
seal, respectively) (Zeisler et al. 1993). These levels corre-
Polar bears
spond well with concentrations found in the western Cana-
1 000
dian Arctic, West Greenland, Svalbard, and northern Nor-
way (Smith and Armstrong 1975, 1978, Johansen et al.
1980, Carlberg and Boler 1985, Wagemann 1989, Wagemann
et al. 1996, Skaare 1994, Dietz et al. 1996, 1997a; Annex
Table 7·A15). Smith and Armstrong (1978) found no signifi-
100
cant differences in methylmercury in liver or muscle tissue
between areas as widely separated as Holman in the western
Canadian Arctic and Pond Inlet on northern Baffin Island.
However, Eaton and Farant (1982) point out, based on the
10
results from Smith and Armstrong (1978), that the age accu-
10
100
1 000
10 000
100 000
mulation of Hg in liver of ringed seals increased from east to
total-Hg, µg/kg ww
west. The increasing trend of Hg in ringed seal liver from
Figure 7·49. Organic mercury versus total mercury in muscle, liver and
eastern to western Canada is supported by a recent study of
kidney tissue of Greenlandic marine animals (after Dietz et al. 1990). The
Wagemann et al. (1996), but no similar trend was found in
lines indicate 100% organic mercury.
kidney and muscle (Annex Table 7·A15 and Figure 7·48).
Polar bears also have quite high Hg levels in their livers.
Juvenile ringed seals from Jarfjord in Norway had lower
Mean concentrations up to 53.0 g/g have been reported in
Hg concentrations in liver and kidney than did ringed seals
polar bear livers from the eastern Beaufort Sea (Norstrom et
from Canadian and Greenland waters. Ringed seals from the
al. 1986, Braune et al. 1991).
Gulf of Bothnia and Gulf of Finland (data not included in
Annex Table 7·A15), on the other hand, exceed even the
Age accumulation. As was the case for Cd, the concentra-
highest values reported for the Arctic (Helle 1981, Perttilä
tion of Hg increases in marine mammals with age. This has
et al. 1986, Frank et al. 1992). Higher levels in seals from
been well documented for seals (Smith and Armstrong 1975,
northwestern European waters than in those from the Arctic
434
AMAP Assessment Report
West Whitefish
Shingle Point
Station
Tuktoyaktuk
60
60
60
45
45
45
30
30
30
15
15
15
0
0
0
Kendall Island
60
45
30
15
0
Paulatuk
60
Hendrickson Island
45
60
30
45
15
30
0
15
0
1993
1994
Arviat
East Whitefish
(Eskimo Point)
60
60
45
45
Repulse Bay
30
60
Grise Fjord
30
15
45
60
15
0
30
45
1993
1994
0
15
30
0
15
East Whitefish Station
Sanikiluaq
0
60
(Belcher Islands)
45
60
30
45
Coral Harbour
15
30
60
Avanersuaq
0
15
45
60
0
30
45
15
30
Umiujaq
0
15
Kuujjuarpik
60
0
(Great Whale)
45
Upernavik
60
30
60
45
15
45
30
Iqaluit
30
0
60
15
15
45
0
0
30
15
0
Kangaatsiaq
60
45
30
15
0
Hg, µg/g ww
Age :
0-1 y
2-4 y
5-15 y
15-15 y
>25 y
Undetermined
Figure 7·51. Distribution of Hg levels in liver tissue of beluga whale (Delphinapterus leucas) of different ages (years). Plots show selected data (geomet-
ric mean values) from Annex Table 7·A15. (Sources of data: see Annex Table 7·A15).
are also indicated by studies involving grey seal (Halichoe-
tion in liver was extremely high, 257-326 g/g (Koeman et
rus grypus) and harbour seal (Phoca vitulina) (Law et al.
al. 1972), most likely due to anthropogenic sources. Yama-
1991, Frank et al. 1992), and in harbour porpoise from the
moto et al. (1987) presented data for two adult Weddell
two areas (Paludan-Müller et al. 1993). In seals found
Sea seals from the more pristine Antarctic. They found
dead off the coast of the Netherlands, the Hg concentra-
0.11-0.16 and 3.1-8.5 g/g Hg in muscle and liver, respec-
Chapter 7 · Heavy Metals
435
Polar bears from northern Alaska have higher levels of
Hg in liver and muscle tissue than do those in western Alas-
ka (Lentfer and Galster 1987). Bears from western Arctic
Eastern
Beaufort Sea
Canada accumulated Hg faster in their livers than did polar
53
Chukchi Sea
4.80
71.1
bears from eastern Arctic Canada, whereas bears from Hud-
38.1
35.9
son Bay had slightly lower concentrations (Norstrom et al.
Amundsen
1986, Braune et al. 1991). The eastward decreasing trend in
Gulf
Hg extended to Greenland and Svalbard (Norheim et al.
23.2
Western
18.3
Beaufort Sea
1992, Dietz et al. 1995; Annex Table 7·A15; Figure 7·52).
S.W.
Spence
22
The observed geographic trend has also been documented
Hadley
Melville Island
Bay
Bay
6.01
for Hg levels in polar bear hair (Eaton and Farant 1982, Ren-
17.8
Northern
13.2
Cornwallis Island
zoni and Norstrom 1990, Born et al. 1991; Figure 7·53).
Hudson Bay
6.54
Southern
Clyde 25.1 Avanersuaq
Norstrom et al. (1986) suggested that the differences found
16.7
Hudson Bay
River
2.6
between western Arctic Canada and eastern Arctic Canada
Cape Mercy
Svalbard
Northern
were most likely caused by higher Hg levels in the ringed
6.72
Baffin Island
9.62
seal food chain caused by higher natural levels in the sedi-
Cumberland
Peninsula
Ittoqqortoormiit
ments (and consequently in the lower food chain) of the
Melville Island area. The evidence presented above indicates
Hg, µg/g ww
a geographic trend in ringed seal liver, but these differences
50
40
are not apparent in kidney and muscle. Too few data are
30
available to evaluate geographical trends in ringed seal blub-
20
10
ber, which is the preferred food of polar bears.
0
Figure 7·52. Distribution of Hg levels (age adjusted to 6.9 years, Svalbard
adults) in liver tissue of polar bear (Ursus maritimus). (Source of data:
Temporal trends. Few investigations on temporal trends
Braune et al. 1991, Norheim et al. 1993, Dietz et al. 1995, 1996).
have been conducted for marine mammals. In a recent study,
18.5
Wagemann et al. (1996) concluded that livers of recently
collected (1987-1993) ringed seals from western Canada
Amundsen Gulf
accumulating Hg approximately three times faster than
Eastern
Beaufort Sea
those collected in 1972-1973 (from Smith and Armstrong
10.2
8.99
1975, 1978; Figure 7·54). In contrast, muscle levels were
significantly lower (43%) in the recent samples, but so was
the mean age (42% lower), which means that muscle levels
1.7
were approximately the same in the two samples. A com-
Wrangel Island
parison of liver values in ringed seals from Avanersuaq,
8.38
northwest Greenland taken in 1984 and 1994 (for age
7.85
1.6
Western
6.59
Cornwallis
Hudson Bay
Lena River
groups: 2-4, 5-10 and > 15 years) showed increasing Hg
6.93
Island
3.0
levels, with concentrations 2.3-6.9 times higher in the re-
Northern
cent samples. Over the same time period, differences were
Baffin
3.0 3.1
Island
minor in one-year old seals from Avanersuaq and Nanor-
2.5
Avanersuaq
3.53
talik, south Greenland. This was true for all age groups from
4.0
Southern Hudson
Ittoqqortoormiit, central East Greenland as well. A recent
4.92
Bay
Ammassalik 4.62
1.98
4.21
Southern
study on Atlantic walruses did not show any temporal trend
Baffin
Svalbard
(Wagemann et al. 1995, 1996), but the authors suggest that
Island
the low Hg levels in walruses and the relatively short sam-
Clyde
Ittoqqortoormiit
River
Mercury
pling time (six years) may explain the lack of trend. The
µg/g ww
same study shows that the Hg accumulation rate in livers of
10
belugas from the western Canadian Arctic increased by a
8
factor of 1.7-1.8 from 1981 to 1993 (Figure 7·54). Kidney
6
levels were also significantly higher (1.74 times), but muscle
4
2
samples were only 25% higher, which was not significant.
0
Total Hg concentrations were also higher in tissues of nar-
Figure 7·53. Distribution of Hg levels in hair of polar bear (Ursus mariti-
whals in 1992-1993 compared with 1978-1979, although
mus). (Sources of data: Eaton and Farant 1982, Renzoni and Norstrom
the size of the animals indicated that the recent group of ani-
1990, Born et al. 1991).
mals was somewhat older (1978-1979, 376 cm; 1992-1994,
tively, which is consistent with background concentrations
420 cm). Pilot whales from the Faeroe Islands have been
in the Arctic.
sampled from 1977 to 1987 (Julshamm et al. 1987, Caurant
The high Hg levels in the western part of the Canadian
et al. 1994). Regressions of Hg concentration in liver against
Arctic have also been documented in beluga whale tissue
time showed no significant trend. To evaluate differences
(muscle, liver, and kidney) and are most clearly seen in liver
over longer periods of time, hair samples from seal fur cloth-
(Annex Table 7·A15 and Figure 7·51). The different ages
ing from the 15th century have been analyzed and compared
of the animals in the two areas were a confounding factor in
with values obtained for recent fur samples, both from Green-
the comparison between the western and the eastern Arctic.
land. The recent values are approximately four times higher
However, comparisons of the accumulation rates of Hg in
(Hansen 1988). When hair concentrations are compared be-
liver (data were not given for muscle and kidney) of belugas
tween Inuit mummies (1470±50 ad) and contemporary
from western and eastern Canada showed that the accumu-
Greenlandic residents, a three-fold increase in Hg concentra-
lation rate was more than three times higher in western Can-
tion is suggested (Hansen 1988). As the Inuit in the 15th cen-
ada (Wagemann et al. 1996).
tury lived almost exclusively on the marine food chain, as
436
AMAP Assessment Report
Hg
shown by the stable isotope 13C/12C ratios, and today's pop-
µg/g ww
ulation lives on a mixed diet of Arctic and imported food,
140
Ringed seal
the increase could have been even larger if compared on a
same-food basis. Analysis results from Ammassalik, East
120
Greenland from the 19th century were all lower than those
from the 15th century, which was attributed to a severe
100
famine in the areas where the samples were taken. The 15th
80
century hair samples are approximately five times higher in
Hg than the present-day Danish level (Hansen 1988).
60
Selenium
40
Selenium analyzes have been completed for most of the spe-
20
cies examined for Hg because Se is regarded as an antago-
nist to Hg, and data interpretation is consequently most val-
0
0
10
20
30
40
50
id when both concentrations are known. In most Arctic sam-
Age
ples, Se is present in a substantial surplus compared with Hg
years
on a molar basis (Dietz et al. 1997a). In ringed seals, Se con-
centrations in liver and kidney are similar, ranging between
Western Arctic 1987-93
Western Arctic 1972-73
Hg = 2.02 + 2.54 Age
Hg = 1.22 + 0.87 Age
0.93 and 18.4 g/g over different age groups and areas. The
levels in muscle are approximately ten times lower than in
liver and kidney. In whales, Se concentrations decrease in
Hg
the order liver > kidney > muscle, whereas in polar bears the
µg/g ww
order is kidney > liver > muscle.
100
Beluga
Mean liver concentrations of Se have been recorded as
high as 18.4 g/g in ringed seals, and the highest levels
80
among the other seven seal species are found in bearded seal
(mean = 34.4 g/g). Walrus livers contain lower Se concentra-
60
tions (highest mean 3.14 g/g) than those of ringed seal. Ba-
leen whales are generally low in Se with a highest mean of
40
1.78 g/g. In toothed whales, the Se level may be as high as
24.2 g/g in belugas. In pilot whales that are known to feed
20
on cephalopods, values are even higher. Samples from 1977
(the year in which high Hg levels were measured) indicate Se
0
concentrations as high as 172 g/g in liver, whereas in other
0
10
20
30
40
50
years they were only 12-23 g/g. Selenium levels up to 23.4
Age
g/g have been recorded in the livers of polar bears.
years
Western Arctic 1993-94
Eastern Arctic 1993-94
Age accumulation. The molar ratio between Se and Hg in
Hg = -15.6 + 2.10 Age
Hg = -0.32 + 0.62 Age
liver tissue of marine mammals is frequently close to 1, at
Western Arctic 1981-84
Eastern Arctic 1984
least at concentrations above approximately 3 g/g (Koe-
Hg = -5.44 + 1.14 Age
Hg = 0.14 + 0.36 Age
man et al. 1973, 1975, Hansen et al. 1990, Braune et al.
1991, Becker et al. 1995a, Dietz et al. 1995). Selenium acts
as a detoxifying agent by binding Hg as mercuric selenide
Hg
(HgSe tiemannite) (Iwata et al. 1981, Joiris et al. 1991).
µg/g ww
It is, therefore, not surprising that Se and Hg accumulate
40
with age in liver tissue. Selenium is generally not found to
accumulate to the same degree with age in muscle and kid-
Beluga
Ringed seal
ney as in liver (Wagemann et al. 1983, 1990, Hansen et al.
1987-93
1990, Paludan-Müller et al. 1993, Dietz et al. 1995). The
30
concentration of Se in the skin of harbour porpoises is posi-
1993-94
tively correlated with age (Paludan-Müller et al. 1993).
As outlined above, Hg and Se are associated in liver. The
1972-73
same geographical trends are therefore found for Se in liver
20
as for Hg, whereas no clear geographical differences can be
detected in other tissues. Hence, Wagemann et al. (1996)
found higher Se levels in ringed seals and belugas from the
1981-84
1993-94
western Canadian Arctic than in those from the eastern
10
Canadian Arctic. This study found no geographical trends
1984
for muscle or skin tissue. Concentrations of Se in marine
mammals have always been found to be higher in northwest
Greenland compared with other Greenland areas, except for
ringed seals, where higher levels in central East Greenland
0
Western Arctic
Eastern Arctic
Western Arctic
have been reported (Dietz et al. 1996).
Figure 7·54. Age accumulation and temporal comparison of mean con-
Temporal trends. No information has been published on
centrations of Hg in liver tissue of ringed seal (Phoca hispida) and beluga
temporal trends for Se. However, it can be expected that the
whale (Delphinapterus leucas) from the Canadian Arctic. (After Wagemann
et al. 1996).
trends found for Hg in liver tissue of ringed seal, belugas,
Chapter 7 · Heavy Metals
437
and narwhals will also be found for Se, since these two ele-
the high concentrations of Cd in the kidneys of some Arctic
ments are highly correlated. Comparisons of ringed seals
mammals (primarily reindeer) and some gamebirds (ptarmi-
from Avanersuaq, northwest Greenland showed an increase
gan). Even though some concentrations exceed threshold
from 1.6 to 4.0 times for the age groups above two years.
values believed to cause kidney dysfunction, no such effects
The younger age groups (0 and 1 years), however, showed a
have ever been observed or reported.
decrease from 0.6 to 0.9 (Annex Table 7·A15).
Cases of severe ecological damage, however, have been
reported in the Arctic terrestrial system, specifically around
the huge metal smelter complexes of the Kola Peninsula, Rus-
7.7. Biological effects
sia. The complete collapse of healthy ecosystems for tens of
(acute, short-, and long-term toxicity;
kilometers around these smelters is primarily the result of pol-
reproductive, physiological,
lution by high concentrations of SO2 and NOx leading to ex-
and behavioral effects; etc.)
treme cases of acidification. Although the concentrations of
metals can be very high in these areas, they are only a contri-
A generalized overview of reported effects threshold levels
buting factor (along with hydrocarbons, anthropogenic chem-
for Hg, Cd, Pb and Se in tissues of main animal groups is
icals, and non-chemical factors) to the damage observed.
presented in Table 7·22. These thresholds, together with in-
formation presented in section 7.5, have been compared
7.7.2. Effects on freshwater ecosystems
with concentrations of metals observed in Arctic biota in the
following assessment of the potential for effects in different
The only biological effect thought to be attributable specifi-
Arctic ecosystems.
cally and unambiguously to heavy metal pollution in the Arc-
tic freshwater ecosystem is the decline in the ringed seal pop-
ulation of the Lake Saimaa, Finland. There it is thought that
7.7.1. Effects on terrestrial ecosystems
Hg contamination (and lack of sufficient Se to detoxify it) has
No biological effects attributable specifically and unambigu-
rendered the seals more prone to premature and still births.
ously to heavy metal pollution in the Arctic terrestrial eco-
It is the view of Russian ecologists that severe ecological
system have been reported. There is concern, however, over
damage has occurred along portions of major Siberian rivers
Table 7·22. Overview of reported effects threshold levels for metals in tissues of main animal groups.
Concentration
Metal
Group
Tissue
mg/kg ww a
Symptoms
Reference
Pb
Waterfowl
Blood
> 20-50
Subclinical poisoning
Pain 1996
> 50-100
Clinical poisoning
> 100
Severe clinical poisoning
Liver
> 2-6
Subclinical poisoning
> 6-15
Clinical poisoning
> 15
Severe clinical poisoning
Bone
> 10-20 (dw)
Subclinical poisoning
> 10-20 (dw)
Clinical poisoning
> 20 (dw)
Severe clinical poisoning
Terrestrial birds
Blood
> 0.2-3
Subclinical poisoning
Franson 1996
> 1-5
Toxic
> 5-10
Lethal
Liver
> 2-6
Subclinical poisoning
> 3-6
Toxic
> 5-20
Lethal
Kidney
> 2-20
Subclinical poisoning
> 3-15
Toxic
> 5-40
Lethal
Mammals
Liver
> 30 (dw)
Clinical signs
Ma 1996
Kidney
> 90 (dw)
Clinical signs
Blood
> 10
Adverse effect on cognitive and behavioral performance
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Cd
Birds
Liver
> 40
Cadmium poisoning
Furness 1996
Kidney
> 100
Cadmium poisoning
Marine mammals
Liver
> 20-200
Potential renal dysfunction
Law 1996
Kidney
> 50-400
Potential renal dysfunction
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Hg
Birds
Liver
> 30
Lethal level in free ranging birds
Thompson 1996
Kidney
> 30
Lethal level in free ranging birds
Egg
> 3.0
Detrimental effect upon free ranging bird hatching
Liver
> 30
Laboratory succumbed animals due to Hg intoxication
Kidney
> 30
Laboratory succumbed animals due to Hg intoxication
Egg
> 2.0
Detrimental effect upon experimental bird hatching
Terrestrial mammals
Liver
> 30
Lethal or harmful in free ranging wildlife
Thompson 1996
Kidney
> 30
Lethal or harmful in free ranging wildlife
Liver
> 25
Laboratory succumbed animals due to Hg intoxication
Kidney
> 25
Laboratory succumbed animals due to Hg intoxication
Marine mammals
Liver
> 60
Liver damage
Law 1996
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
Se
Birds
Eggs
> 3
Deformed embryos
Heinz 1996
Liver
> 9
Deformed embryos
Kidney
> 9
Deformed embryos
Blood
> 5-14
Lethal effect
Terrestrial mammals
Liver
> 7
Hepatic lesions
WHO 1997
a. Unless indicated as dw.
438
AMAP Assessment Report
Cd
Pb
µg/g ww
µg/g ww
10
200
9
8
7
6
100
5
90
80
4
70
60
3
50
40
2
30
20
10.9
0.8
0.7
0.6
10
9
0.5
8
7
0.4
6
0.3
5
4
0.2
3
2
0.1
0.09
0.08
0.07
1
0.06
0.9
0.05
0.8
0.7
0.04
0.6
0.5
0.03
0.4
0.02
0.3
0.2
0.01
0.009
0.008
0.007
0.1
0.006
0.09
0.005
0.08
0.07
0.004
0.06
0.05
0.003
0.04
0.002
0.03
0.02
0.001
Muscle
Liver
Kidney
0.01
Muscle
Liver
Kidney
Fish
Seal
Fish
Seal
Seal
Whale
Whale
Whale
Seabird
Seabird
Seabird
Polar bear
Polar bear
Fish
Fish
Fish
Seals
Seals
Seals
Prawn (meat)
Whales
Whales
Whales
Prawn (shells)
Seabirds
Seabirds
Seabirds
Decapods
Polar bear
Polar bear
Polar bear
Capelin (whole fish)
Molluscs (soft tissue)
Molluscs (soft tissue)
Figure 7·55. Summary of ranges of Pb levels (mean values) in Arctic ma-
Copepods and other zooplankton
Decapods (heads, shells or whole)
rine organisms. Solid parts of the lines indicate ranges for Greenlandic
data from Dietz et al. (1996), where the Pb analyses have been critically
evaluated and only data for samples with n 8 were included.
Figure 7·56. Summary of ranges of Cd levels (mean values) in Arctic ma-
and watercourses. As noted above for the terrestrial ecosys-
rine organisms. Solid parts of the lines indicate ranges for Greenlandic
data from Dietz et al. (1996), where the Cd analyses have been critically
tems, however, it is difficult to conclude that metal contamina-
evaluated and only data for samples with n 8 were included.
tion is any more than one of many factors which have caused
the problem. In many of the most-affected parts of the aquatic
tissue burdens, potential effects cannot be derived from the
ecosystem, large inputs of sewage, petroleum hydrocarbons,
levels measured. The only places where effects are likely to
sulfactants, refined petroleum products, acidifying compounds,
occur are hot spots such as mining areas and possibly some
and other anthropogenic chemicals are common. These sub-
Russian estuaries.
stances are capable of causing the damage even without the
additional negative consequences of metal pollution.
Cadmium
According to the information from the literature summa-
rized in section 7.5.2, effects on molting might occur in
7.7.3. Effects on marine ecosystems
shrimps when body burdens exceed 10-40 g/g. However,
Lead
Cd in Arctic invertebrates is lower than this limit and there-
The overall Pb baseline levels in the Arctic are low. There is
fore no effects are expected in this animal group. No infor-
no indication that Pb levels increase in higher trophic levels.
mation relating biological effects to body concentrations is
As lethal and sublethal effects of Pb have not been related to
available for fish; consequently, effects can not be evaluated
Chapter 7 · Heavy Metals
439
Hg
Se
µg/g ww
µg/g ww
300
200
200
100
90
100
80
90
70
80
60
70
60
50
50
40
40
30
30
20
20
10
10
9
9
8
8
7
7
6
6
5
5
4
4
3
3
2
2
1
1
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.5
0.5
0.4
0.4
0.3
0.3
0.2
0.2
0.1
0.09
0.1
0.08
0.09
0.07
0.08
0.06
0.07
0.05
0.06
0.05
0.04
0.04
0.03
0.03
0.02
0.02
0.01
Muscle
Liver
Kidney
0.01
Muscle
Liver
Kidney
Fish
Fish
Fish
Seals
Seals
Seals
Whales
Whales
Whales
Seabirds
Seabirds
Seabirds
Fish
Fish
Fish
Seals
Seals
Seals
Polar bear
Polar bear
Polar bear
Amphipods
Whales
Whales
Whales
Seabirds
Seabirds
Seabirds
Polar bear
Polar bear
Polar bear
Amphipods
Molluscs (soft tissue)
Molluscs (soft tissue)
Copepods and
Copepods and
Decapods (whole animals)
Crustaceans (whole animals)
Figure 7·57. Summary of ranges of Hg levels (mean values) in Arctic ma-
rine organisms. Solid parts of the lines indicate ranges for Greenlandic
data from Dietz et al. (1996), where the Hg analyses have been critically
evaluated and only data for samples with n 8 were included.
Figure 7·58. Summary of ranges of Se levels (mean values) in Arctic ma-
rine organisms. Solid parts of the lines indicate ranges for Greenlandic
for this animal group. Kidney damage can occur in pelagic
data from Dietz et al. (1996), where the Se analyses have been critically
sea birds at Cd concentrations of 60 g/g, whereas the cor-
evaluated and only data for samples with n 8 were included.
responding limit appears to be approximately 100 g/g for
marine mammals. The frequencies of occurrence of Cd con-
cous gulls from Avanersuaq and kittiwakes from Upernavik,
centrations above these limits were calculated for marine
as many as 50% of the birds contained Cd exceeding the 60
birds and seals from the Greenland area, where raw data
g/g limit.
were available and geographical comparisons indicated the
A similar pattern was observed for seals in Greenland wa-
highest Cd levels. In no cases were kidney levels above 60
ters. The limit of 100 g/g was exceeded in 10.3% of the
g/g for seabirds at ages 0 and 1 year. For birds older than
ringed seals and in 5% of hooded seals (Cystophora cristata)
one year, 7.3% of the birds had concentrations higher than
analyzed, whereas no values above this threshold were re-
60 g/g (see also Figure 7·59, next page). This percentage
ported for harp seals. Again the high values seem to be found
increased northward with 7.9, 13.3, and 18.8% in Uum-
in northwest Greenland, where as many as 37.0 and 35.2%
mannaq, Upernavik, and Avanersuaq, respectively. For glau-
of the ringed seal measurements from Avanersuaq and Uper-
440
AMAP Assessment Report
Cd
Cd
µg/g ww
µg/g ww
200
Black guillemot
600
Brunnichs guillemot
Danmarkshavn
180
Common eider
Ittoqqortoormiit
Glaucous gull
Kong Oscars Fjord
160
King eider
500
Nanortalik
Longtailed duck
Iceland gull
Avernersuaq
140
Redbreasted merganser
Upernavik
400
Ivory gull
120
Uummannaq
Kittywake
Svalbard
Northern fulmar
100
300
80
200
60
40
100
20
0
0 0
5
10
15
20
25
30
35
40
Avanersuaq
Upernavik
Uummannaq
Kangaatsiaq
Nanortalik
Itoqqortormit
Age
Figure 7·59. Cadmium levels in kidney of adult marine birds from Green-
Figure 7·60. Cadmium levels in kidney of adult ringed seals from Green-
land. Levels associated with potential for kidney damage are above 100
land and Svalbard. Levels associated with potential for kidney damage are
µg/g ww. (Source of data: R. Dietz unpubl.).
above 100-200 µg/g ww. (After Dietz et al. 1998 in press).
navik, respectively, were above the 100 g/g limit (see Fig-
and significant decreases in intestinal absorption. These ef-
ure 7·60). In the human health chapter (chapter 12) the au-
fects of chronic exposure to Hg may appear at tissue bur-
thors claim that effects may be found in humans even at 50
dens above 25-60 µg/g ww (see Table 7·22).
g/g as the `healthy worker effect' may cause threshold lim-
its to be set too high. In the 10-15 year-age-group, as many
Selenium
as 66.7% of the seals from Avanersuaq exceeded the 100
Few studies of the effects of chronic exposure to Se have
g/g limit, as did 50% of the 5-10 year-age-group from both
been related to the respective tissue burdens. In a previously
Avanersuaq and Upernavik. In order to determine whether
cited experiment where hepatic lesions were found in rats
these apparently high levels pose a problem to the seals,
fed on an Se-rich diet, the liver Se concentration was 7.34
samples of kidneys from this area in Greenland were exam-
g/g. A substantial proportion of Arctic marine biota have
ined in a pilot study by skilled pathologists. Fiftheen kidney
liver concentrations above this limit, but Se in the marine
samples were selected from three different concentration
ecosystem in general is higher than in the terrestrial ecosys-
ranges (< 2, 80-100, and > 200 g/g ww) for histopatholo-
tem. The observed 1:1 molar ratio between Se and Hg com-
gical examination. The examination did not reveal any kid-
monly found in liver tissue of marine mammals suggests that
ney damage even in the highest concentration group, sug-
these metals act as antagonists for each other and are detox-
gesting that seals have an ability to detoxify substantial
ified as inert mercuric selenite (HgSe, or tiemannite).
quantities of Cd (Dietz et al. 1998 in press).
In conclusion, Cd concentrations in kidneys of some spe-
cies of marine birds and mammals, especially in northwest
7.8.
Greenland, are high enough to cause concern. Such high
Conclusions and recommendations
values may be seen in other Arctic areas as well. As only
7.8.1. Conclusions
one recent pilot study has focused on possible biological
Details are given in Tables 7·23 and 7·24.
effects related to high Cd tissue burdens (such as morpho-
logical damage at the microscopic level), it is at present im-
1. The distribution of heavy metals among the various envi-
possible to make firm conclusions as to whether these ani-
ronmental compartments of the Arctic is dynamic and dri-
mals are suffering from any effects of the high Cd levels.
ven by natural sources, processes, and environmental fac-
A preliminary conclusion of the pilot study is that Arctic
tors. Significant anthropogenic inputs of metals are detect-
seal populations may have developed strategies to effectively
able against the highly variable natural background on local
detoxify Cd.
scales, commonly in the order of tens of kilometers or less.
2. Metals are taken up by Arctic biota and their levels often
Mercury
reflect local geology or local anthropogenic activities.
As no studies of the effects of chronic exposure to Hg have
3. The most important metals in the Arctic biosphere are Cd
been related to tissue burdens, it is not possible to determine
and Hg because they occur in some biota at concentra-
what tissue levels constitute a risk in marine biota. Experi-
tions that may have health implications for individual an-
ments on certain animal groups have shown that the central
imals or may have implications for human consumers.
nervous system and kidney are the organs most susceptible
4. Few spatial or temporal trends are apparent in the exist-
to damage from methylmercury and inorganic Hg, respec-
ing data, largely due to poor temporal or spatial coverage
tively. Even for humans, no guideline concentration limits
or to unresolvable artifacts in the data related to differ-
are available for internal organs because the parameter most
ences in sampling, analytical, and reporting protocols.
widely used to measure Hg burden is blood concentration.
It would therefore be relevant to carry out morphological
7.8.1.1. Sources and transport of metals
effect studies in populations known to have high body bur-
dens of Hg, for example animals from the western Canadian
1. During winter, about two-thirds of the heavy metals in air
Arctic. Physiological and biochemical abnormalities include
in the High Arctic are transported from Eurasia, particu-
neurological impairment, reproductive effects, liver damage,
larly from the Kola Peninsula, the Norilsk region, the
Chapter 7 · Heavy Metals
441
Table 7·23. Overview of major conclusions related to terrestrial/aquatic ecosystems.
Air and
atmospheric
deposition
Freshwater
Sediment
Soil
Vegetation
Birds
Fish
Mammals
1. Concentrations of metals exceeding average global background
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
1.1. Regional None
None
None
None
None
High Cd in
None
High Cd in
kidney of
reindeer kidney
ptarmigan
1.2. Local
Kola Peninsula At `hot spots'
At `hot spots'
At `hot spots'
At `hot spots'
Cd in kidney
High Hg
Unknown
of human/indus- of human/indus- of human/indus- of human/indus- of ptarmigan
in fish of
trial activity
trial activity
trial activity
trial activity
high in
NWT
such as smelter
such as smelter
such as smelter such as smelter Yukon/NWT
complexes of
complexes of
complexes of
complexes of
the Kola Penin- the Kola Penin- the Kola Penin- the Kola Penin-
sula (scale 10-
sula (scale 10-
sula (scale 10-
sula (scale 10-
100 km)
100 km)
100 km)
100 km)
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
2. Spatial trend
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
2.1. Regional South to north
None
High Hg
Kola/Northern
Kola/Northern None
None
None
decrease
concentrations
Scandinavia Scandinavia
in Arctic lakes
enrichment
enrichment
in Cu, Ni
in Cu, Ni
2.2. Local
Decrease with
Enrichment near Gradients near
Gradients near
Gradients near Cd in
Highest
Cd in caribou
distance from
point sources
point sources
point sources
point sources
ptarmigan
values for
kidney higher
the source
(Russian rivers, (Cu, Ni)
(Cu, Ni)
(Cu, Ni)
high in
Hg occur
in Yukon
region
lakes of Kola
Yukon
in Canada
than in NWT
Peninsula)
Pb in herbivores
In Norway Cd
in Russia
in reindeer/
greater in east
moose kidney
than in west
increases along
N-S gradient
In Russia Pb in
reindeer liver/
muscle higher in
east than in west
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
3. Temporal trends
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
3.1. Regional A decrease over Unknown
Recent Arctic-
Unknown
Unknown
Unknown
Unknown
Unknown
last 2 decades.
wide increase
Strong seasonal
in surficial Hg
variation; the
concentrations
highest values
in Arctic lakes
seen in winter
3.2. Local
A decrease over Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
the last 5 years
in the Kola Pen-
insula. Winter
concentrations
higher than
in summer
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
4. Observed biological effects attributable to metals
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
4.1. Regional Not applicable Not applicable
Not applicable
Not applicable
Combined
Combined
Combined
Combined
4.2. Local
Not applicable Not applicable
Not applicable
Not applicable
Combined
Cd in some
Combined
Cd in some
ptarmigans
moose and rein-
from Yukon is
deer from Yu-
high enough
kon are high
to cause kidney
enough to cause
damage
kidney damage.
Possibly still
births of Lake
Saimaa ringed
seal due to Hg
contamination.
Unknown: Insufficient data to reach a conclusion.
None: No trend or effect documented from a fair amount of data.
Combined: Contamination by metals may, and probably does contribute to some effects caused primarily by other factors.
Urals, and the Pechora Basin. Five to ten percent of these
the Kola Peninsula and at Norilsk and result from emis-
emissions are deposited in the High Arctic. The remaining
sions from these smelters.
one third of the heavy metals in High Arctic air in winter
3. Near point sources such as mine sites and some Russian
is transported from industrial regions in Europe and
estuaries, heavy metals exceed background levels up to 30
North America. In summer, local sources dominate the
km from the source.
contamination of the High Arctic.
4. Riverine transport of heavy metals toward the Arctic Ba-
2. The highest concentrations of atmospheric heavy metals
sin is approximately half the atmospheric contribution for
in Arctic air occur in the vicinity of smelter complexes on
metals like Cd and Pb, while for others such as Zn the riv-
442
AMAP Assessment Report
Table 7·24. Overview of major conclusions related to the marine ecosystem.
Seawater
Sediment
Algae
Invertebrates
Fish
Seabirds
Marine mammals
1. Concentrations of metals exceeding average global background
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
1.1. Regional None
None
None
Cadmium is high- None
None
Cadmium is higher
er in Arctic mus-
i
n some Arctic
sels and prawns
marine mammals
1.2. Local
At point sources
At point sources
At point sources
At point sources
At point sources
None
None
such as mining
such as mining
such as mining
such as mining
such as mining
areas in Canada
areas in Canada
areas in Canada
areas in Canada
areas in Canada
and Greenland
and Greenland
and Greenland
and Greenland
and Greenland
as well as some
as well as some
as well as some
as well as some
as well as some
Russian estuaries Russian estuaries
Russian estuaries Russian estuaries
Russian estuaries
(scale <30 km)
(scale <30 km)
(scale <30 km)
(scale <30 km)
(scale <30 km)
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
2. Spatial trends within the Arctic
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
2.1. Regional None
Sediments seem
None
Cadmium is
Arctic cod from
Seabirds from
The highest Cd
to follow
higher in
the Barents Sea
Lancaster Sound
levels are recorded
geographical
Greenland
are lower in Cd
have high Cd
in northwest Green-
provinces over
compared
than the rest of
levels, whereas
land and the lowest
the Arctic
to Norway
the Arctic. The
birds from the
are from Svalbard.
highest level
Svalbard region
Hg levels are high-
for Cd and Hg
have low
est in western
are seen at several
Canada and in pilot
populations from
whales from the
northern latitudes
Faeroe Islands
in Canada and
W Greenland. Hg
is low in fish from
the Greenland Sea
2.2. Local
Enrichment near
Enrichment near
Enrichment near
Enrichment near
Enrichment near
Not likely,
Not likely,
point sources. In- point sources. In- point sources.
point sources.
point sources.
due to the migra- as few marine mam-
creasing natural
creasing natural
Increasing natural Increasing natural Increasing natural tory behavior
mals are stationary
Cd from inner
Cd from inner
Cd in individuals Cd in individuals
Cd in stationary
of birds
fjords toward
fjords toward
from inner fjords from inner fjords
fish from inner
the sea
the sea
toward the sea
toward the sea
fjords toward
the sea
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
3. Temporal trends within the Arctic
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
3.1. Regional Higher Pb
Hg in sediment
Unknown
Unknown
Unknown
Moderate to no
Hg in seals from
levels than in
from most Arctic
increase in Hg
northern Canada
prehistoric
areas show levels
and Greenland
time
elevated or
as well as
increasing in
toothed whales
recent sediments
are increasing
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
4. Observed biological effects attributable to metals
-------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------
4.1. Regional Not applicable
Not applicable
Not likely
Not likely
Not likely
Cadmium is high Cadmium is high
enough in some
enough in some
areas to pose a
areas to pose a
threat for kidney
threat for kidney
damage
damage
4.2. Local
Not applicable
Not applicable
Possible
Possible
Possible
Cd is high enough Cd is high enough
combined effect
combined effect
combined effect
in some areas to
in some areas to
in some Russian
in some Russian
in some Russian
pose a threat for
pose a threat for
estuaries
estuaries
estuaries
kidney damage
kidney damage
Unknown: Insufficient data to reach a conclusion.
None: No trend or effect documented from a fair amount of data.
Combined: Contamination by metals may, and probably does contribute to some effects caused primarily by other factors.
ers are more important, carrying five times the atmos-
sured on the Kola Peninsula are comparable with the con-
pheric load. Such mass balance calculations will change
centrations in the most polluted regions of Europe and
considerably with the distance from the sources and the
North America.
time of year, since the source contributions are strongly
2. Background levels in soil, lakes, rivers, and oceans gener-
seasonal.
ally fall within the global ranges.
3. Cadmium levels in some terrestrial birds and mammals
are high compared with global background, as are Hg
7.8.1.2. Arctic metal concentrations
levels in some freshwater fish. Cd levels in marine organ-
relative to global background
isms from large parts of the Arctic exceed global back-
1. Heavy metal concentrations in air in the High Arctic are
ground. Mercury and Se levels in marine mammals are
one order of magnitude lower than concentrations in
high, but do not exceed the highest global levels. Lead
other remote locations and about two orders of magni-
levels in large parts of the Arctic are at the lower end of
tude lower than the concentrations around major point
global background.
sources in the Kola Peninsula. Air concentrations mea-
Chapter 7 · Heavy Metals
443
7.8.1.3. Spatial trends within the Arctic
7.8.1.5.2. Tissue burdens of metals
relative to national standards
1. The concentrations in surface deposition around the
sources, e.g., on the Kola Peninsula, decrease by between
1. Relative to the concentration limits proposed by the Nor-
one and two orders of magnitude within 10-100 km from
dic Council of Ministers for Cd in kidney, liver, and mus-
the emission source. The concentrations within the large
cle tissue, all caribou in Canada as well as most game
area of the High Arctic are uniformly distributed, varying
birds and marine mammals across the Arctic have exces-
by a factor of 2-3.
sive levels of Cd. In almost all cases, Pb levels in marine
2. The concentrations of trace elements in marine sediments
organisms from the Arctic are well below food standard
are dependent on local geology, particle size, the amount
limits; however, this is not the case for hot spot areas such
of organic matter, and anthropogenic influence. The back-
as mining areas and some Russian estuaries. No standard
ground geographical distribution of Pb, Cd, Hg, and Cu
limits are given for Se in food, but in some cases, human
in marine sediments is related to the geological provinces
intake of Se is estimated to be high.
of the Arctic.
3. Regional geographical differences in metal concentrations
7.8.2. Recommendations
of benthic flora and fauna as well as in those of fish are
not very apparent. Seasonal and local trends for a metal
1. No additional heavy metal data should be collected for
like Cd for some stationary marine species are larger than
AMAP applications until unified, standard protocols for
regional differences in baseline data. No geographical dif-
sampling, analysis, and reporting are established for the
ferences can be observed in fish species. For marine birds,
AMAP countries or unless it can be clearly shown in ad-
only Cd shows a geographical trend, seeming to be high-
vance that data to be obtained are sufficiently tied, through
est in northwest Greenland and the Lancaster Sound area.
a comprehensive QA/QC protocol, to coherent standards
For ringed seals, beluga whales, and polar bears, Cd lev-
or methods that ensure final intercomparability of datasets.
els have been shown to be highest in eastern Arctic Can-
2. Mass balance studies for heavy metals in the air, rivers,
ada and northwest Greenland. Mercury levels for the
and ocean across the Arctic should be conducted. Increased
same group of animals have proven to be highest in the
source receptor modeling is required to develop quantita-
western Canadian Arctic, decreasing toward south and
tive strategies for the basis of formulating policy decisions
east. Geology, food constitution, and growth processes
among states.
linked to temperature are possible explanations for these
3. The data gaps for biota should be filled, but priority
differences.
should be given to metals and organisms for which there
are concerns for biological effects. The lack of data from
e.g. marine mammals of Russian waters should be addressed.
7.8.1.4. Temporal trends within the Arctic
4. Studies on the processes behind the geographical differ-
1. The concentration of heavy metals measured in subarctic
ences and trends observed in the Arctic should be initi-
air has decreased during the last two decades. All the
ated and supported.
heavy metals show strong seasonal variation in the High
5. Additional data should be acquired to elucidate Cd and Hg
Arctic.
time trends in a wide number of Arctic ecosystem compart-
2. Mercury in Arctic sediments shows an increase over time,
ments. Processes behind potential trends should be studied
indicating a widespread regional process. As the anthro-
to resolve whether the changes are natural or man-made.
pogenic fluxes does not show the same pattern, further
6. Health effects should be studied in Arctic species having
investigations are needed before firm conclusions can be
body burdens containing Cd and Hg levels of concern.
drawn.
For recommendations on tissue burdens of metals relative to
3. Temporal trend data are scarce in Arctic biota. There is
national standards see chapter 12 (Human Health).
some evidence of Hg increasing by a factor 2-3 in some
marine mammals within the last two decades. Only liver,
and in certain cases kidney, shows such increase. It re-
Acknowledgments
mains uncertain, however, whether this is a real increase
or reflects year-to-year variation. Mercury concentrations
Editors
in human and seal hair from the 15th century are 2-3
Rune Dietz, Jozef Pacyna, David J. Thomas.
times lower than present-day samples.
Authors
Rune Dietz, Jozef Pacyna, David J. Thomas, Gert Asmund,
7.8.1.5. Observed biological effects and health aspects
Viacheslav V. Gordeev, Poul Johansen, Vitaly Kimstach, Lyle
attributable to metals
Lockhart, Stephanie Pfirman, Frank Riget, Glen Shaw, Rudi
7.8.1.5.1. Observed biological effects
Wagemann, Mark White.
1. Health effects have so far not been investigated in Arctic
Unpublished data contributors
biota. However, Cd levels in some caribou, moose, and
L. Barrie, C. Bezte, E.W. Born, B. Braune, Canadian Wildlife
ptarmigan from the Yukon Territory as well as those in
Service, B. Fallis, J. Ford, Freshwater Institute (Winnipeg,
seabirds and marine mammals from northwest Greenland
Canada), M. Gobeil, C. Hendzel, Institute of Nutrition, In-
may be high enough to cause kidney damage. It is like-
ternational Council for Exploration of the Sea, M. Kingsley,
wise uncertain whether Hg poses a health threat to the
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Contents
Whoriskey, F.G. and D.T. Brown, 1988. Mercury levels in lake herring
(Coregonus artedii) and rock cod (Gadus ogac) from the Wemindji
Table 7·A1. Metals in soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 454
region of James Bay. Department of Renewable Resources, MacDonald
Table 7·A2. Metals in vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . 455
Coolege of McGill University, Rapport à la direction Environment
Table 7·A3. Metals in terrestrial/aquatic birds . . . . . . . . . . . . . . . . . 459
Hydro-Québec, 78p.
Table 7·A4. Metals in terrestrial mammals . . . . . . . . . . . . . . . . . . . 462
Windom, H.L., S.J. Schropp, F.D. Calder, R.J. Smith Jr. L.C. Burney, F.G.
Table 7·A5. Metals in freshwater sediment . . . . . . . . . . . . . . . . . . . 467
Lewis and C.H. Robinson, 1989. Natural trace metal cocentration in
Table 7·A6. Metals in freshwater particulates . . . . . . . . . . . . . . . . . 468
estaurine and coastal marine sediments of the southeastern U.S. Envi-
Table 7·A7. Metals in freshwater. . . . . . . . . . . . . . . . . . . . . . . . . . . 469
ron. Sci. Technol. 23: 314-320.
Table 7·A8. Metals in freshwater invertebrates . . . . . . . . . . . . . . . . 470
Wong, M.P., 1985. Chemical residues in fish and wildlife species harvested
Table 7·A9. Metals in freshwater fish. . . . . . . . . . . . . . . . . . . . . . . . 470
in Northern Canada. Environmental Studies Program, Northern Envi-
Table 7·A10. Metals in marine sediments . . . . . . . . . . . . . . . . . . . . . 473
ronment Directorate, Indian and Northern Affairs Canada, pp. 5.10-
Table 7·A11. Metals in marine algae . . . . . . . . . . . . . . . . . . . . . . . . . 481
6.5.
Table 7·A12. Metals in marine invertebrates . . . . . . . . . . . . . . . . . . . 483
Wren, C.D., T. Nygård and E. Steinnes, 1994. Willow ptarmigan (Lago-
Table 7·A13. Metals in marine fish . . . . . . . . . . . . . . . . . . . . . . . . . . 487
pus lagopus) as a biomonitor of environmental metal levels in Norway.
Table 7·A14. Metals in marine birds . . . . . . . . . . . . . . . . . . . . . . . . . 495
Environ. Pollut. 85: 291-295.
Table 7·A15. Metals in marine mammals . . . . . . . . . . . . . . . . . . . . . 504
Yamamoto, Y., K. Honda, H. Hidaka and R. Tatsukawa 1987. Tissue dis-
Table 7·A16. Metals in wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . 522
tribution of heavy metals in Weddell seals (Leptonychotes weddellii).
Table 7·A17. Guideline values . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 524
Document Outline
- Go to opening screen
- 7.0. Introduction
- 7.1. Physical /chemical characteristics
- 7.1.1. Identification of metals to be considered
- 7.1.2. Speciation of metals in the environment
- 7.2. Sources of pollution
- 7.2.1. Natural sources
- 7.2.2. Anthropogenic sources
- 7.2.3. Emission inventories for sources within and outside the Arctic
- 7.3. Special issues of pathways of metals was estimated that 8-16 tonnes of Zn, 6-12 tonnes of Pb,
- 7.3.1. Atmospheric transport
- 7.3.2. Rivers
- 7.3.3. Estuaries
- 7.3.4. Oceans
- 7.3.5. Ice
- 7.4. Toxicological characteristics
- 7.4.1. Toxicokinetics: general principles
- 7.4.2. Uptake
- 7.4.3. Transport, biotransformation, and distribution
- 7.4.4. Excretion
- 7.4.5. Uptake, accumulation, and loss in biota
- 7.5. Toxicological effects
- 7.5.1. Lead
- 7.5.2. Cadmium
- 7.5.3. Mercury
- 7.5.4. Selenium
- 7.6. Regional and circumpolar levels and trends of metal contaminants
- 7.6.1. Atmosphere
- 7.6.2. Terrestrial ecosystems
- 7.6.3. Freshwater ecosystems
- 7.6.4. Wetland ecosystems
- 7.6.5. Marine ecosystem
- 7.7. Biological effects (acute, short-, and long-term toxicity; reproductive, physiological, and behavioral effects; etc.)
- 7.7.1. Effects on terrestrial ecosystems
- 7.7.2. Effects on freshwater ecosystems
- 7.7.3. Effects on marine ecosystems
- 7.8. Conclusions and recommendations
- 7.8.1. Conclusions
- 7.8.2. Recommendations
- Acknowledgments
- References
- Annex