ICES Journal of Marine Science, 55: 34–43. 1998
Tributyltin inputs to the North Sea from shipping activities, and
potential risk of biological effects
Ian M. Davies, Susan K. Bailey, and
Melanie J. C. Harding
Davies, I. M., Bailey, S. K., and Harding, M. J. C. 1998. Tributyltin inputs to the
North Sea from shipping activities, and potential risk of biological effects. – ICES
Journal of Marine Science, 55: 34–43.
The mechanisms leading to an annual input of tributyltin (TBT) to the North Sea from
shipping are discussed. It is estimated that the gross annual input of TBT to the North
Sea was 68 tonnes.
A simple numerical model of the North Sea has been applied to the input data, and
indicates that the greatest risk of biological effects of TBT from commercial vessels
would be expected in areas of the south-eastern North Sea (NSTF Areas 4, 5), and off
eastern England (Area 3ii). Low intensity of impact was indicated in the northern
(Areas 1, 2, 3i) and north-central (Area 7i) North Sea. The sensitivity of the model
output to variations in critical parameters of the behaviour and inputs of TBT are
discussed, and are found to have minor effects on the overall conclusions.
The estimates of relative risk of biological effects from TBT broadly reflect the
impacts observed in dogwhelks (Nucella lapillus L.) in North Sea coastal areas.
? 1998 International Council for the Exploration of the Sea
Key words: tributyltin, marine modelling, inputs, effects, dogwhelk, Nucella.
Received 8 December 1995; accepted 17 January 1997.
I. M. Davies, S. K. Bailey, and M. J. C. Harding: SOAEFD Marine Laboratory,
PO Box 101, Victoria Road, Aberdeen, AB11 9DB, Scotland, UK. Correspondence to
I. M. Davies: fax: +44 1224 295511; email: [email protected].
Introduction
Since the introduction by several North Sea states of
statutory controls on the use of tributyltin (TBT) compounds in antifoulants on small vessels and in mariculture, the major remaining use of TBT in the sea has
been for the protection of large vessels. Field studies
around harbours (e.g. Bailey and Davies, 1988) and
major shipping routes (e.g. Ten Hallers-Tjabbes et al.,
1994) have indicated that TBT from shipping can cause
significant biological effects in sensitive species.
TBT compounds from shipping may enter the
environment in five ways:
(a) at the time of ship construction,
(b) when vessels are stationary in port (static
leaching),
(c) when vessels are under way at sea (dynamic
leaching),
(d) when vessels are in dry dock for repair and
maintenance, including repainting,
(e) when vessels are scrapped.
These inputs will be unevenly distributed over the
North Sea, and will be subject to differing degrees of
1054–3139/98/0100034+10 $25.00/0/jm970275
dispersion and dilution according to the hydrographic
characteristics of various areas. For example, in the
middle of the North Sea the main inputs will be from
vessels on passage; however, the receiving waters in this
area are slow moving. The faster moving water in the
southern North Sea and the Channel has more traffic
and harbours than other parts of the North Sea, and
might receive greater inputs of TBT. The interactions
between TBT and water fluxes may therefore lead to
different intensities of effects on organisms around the
coasts of the North Sea.
To provide a framework to assist in the interpretation
of a study of the impact of TBT on dogwhelks on the
coast of the North Sea (Harding et al., 1992), an
estimate has been made of the releases of TBT at that
time from large vessels in various phases of their operation (on passage, in port, undergoing refit, etc.), as
sources of TBT to the North Sea. These estimated inputs
have been partitioned geographically between sea and
coastal areas, and incorporated in a box model of the
circulation of the North Sea to provide indications of
the relative risk in these areas of biological effects from
TBT from shipping. The sensitivity of the estimates of
? 1998 International Council for the Exploration of the Sea
Tributyltin inputs to the North Sea
35
Table 1. Maximum inputs of TBT to the North Sea from individual states, based on the percentage of
world trade entering each state (Stopford, 1988; A Milne, pers. comm.).
State
UK
France
Belgium
Netherlands
Germany
Denmark
Finland
Sweden
Norway
Imports
(M t)
Percentage
of world
trade
TBT
input
(t)
101
94
66
229
81
30
31
48
16
2.91
2.71
1.90
6.60
2.33
0.86
0.89
1.38
0.46
34.93
32.51
22.82
79.19
28.01
10.37
10.72
16.60
5.53
696
20.04
240.68
relative degrees of contamination to the assumptions
inherent in the model is explored.
The outputs from the box model are expressed as a
ranking of the risk of biological effects in various sea
areas. Little weight is given to the absolute values of the
estimates of concentration of TBT in the sea water, as
some of the environmental processes affecting TBT
concentrations (e.g. interactions with solids) are not
included in the model. However, the estimates are used
to give an indication of relative risk of effects, for
comparison with results of a broad-scale survey of
imposex in dogwhelks (Harding et al., 1992).
Estimation of the input of TBT to the
North Sea from commercial shipping
Global input of TBT resulting from shipping
activities
A first approach to estimating the annual input of TBT
to the North Sea from commercial shipping may be
made from consideration of the total global usage of
TBT, and the proportion of global seaborne trade that
passed through the North Sea. The amount of TBT in
paint produced each year is taken as an estimate of the
annual quantity lost from shipping. The total quantity
of marine paint produced annually was approximately
25.106 l, of which about 80% was TBT-based (A. Milne,
pers. comm.), i.e. approximately 20.106 l of TBT based
paint produced annually. As the specific gravity of paint
is about 1.5 g cm "3, this equates to approximately
30 000 t of paint. The average TBT content of marine
paints was about 4%, and therefore approximately
1200 t of TBT was applied annually to ships’ hulls.
In 1984, the volume of global seaborne imports was
3470 t#106 t, of which 20% passed through the North
Sea (Stopford, 1988). On this basis, the maximum
amount of TBT that might be released into the North
Comment
82.7% of total of 122 M t of imports
50% of total of 197 M t of imports
Assuming leakage from Baltic
As above
Sea from shipping may be estimated as 240 t (20% of
1200 t), and this figure may be roughly partitioned
between states in proportion to their volumes of maritime trade (Table 1). This estimate of 240 t must be
considered as a maximum value, against which to view
the more detailed assessment below.
Estimates of TBT inputs from specific sources
A more detailed estimate of the input of TBT from
shipping to the North Sea can be obtained by consideration of the various input mechanisms individually.
The total input estimated by this method should not
exceed 240 t.
From ship building and dry-docking
The annual production of paints containing 1200 t of
TBT will be used on the hulls of new vessels, and on
older vessels during dry-docking. The data for the
following calculations were obtained from Stopford
(1988).
(i) From ship building. 630 million (M) dead weight
tonnage (DWT) of shipping was used worldwide. The
life expectancy of these ships was about 25 yr, so to
replace the lost tonnage required 25 M DWT of shipping
to be built each year. The majority (70%) was built in
Japan and Korea, and only approximately 15% in the
whole of Europe. The only significant builders along the
North Sea coast were in Denmark and Germany, which
accounted for around 25–30% of the total ship building
in Europe.
Typical periods between dry-docking were 2–2.5 yr, so
250 M DWT of shipping were dry-docked each year.
New building (25 M DWT) was therefore about 10% of
the total dock activity, i.e. used 120 t of TBT. It is
estimated that perhaps 1% might be expected to be lost
to the environment in solution (i.e. 1.2 t). The maximum
36
I. M. Davies et al.
amount of TBT lost in all European waters was
therefore 15% of this (i.e. 0.18 t), 0.05 t of which would
be in the North Sea.
(ii) Losses from dry docking activities. Each year, 25 M
DWT of shipping (10% of the total trade) were drydocked in Northern Europe. Loss of TBT during dry
docking will arise from hull washings/scrapings, and the
over-spray of new paint material. The latter is considered to be a very small proportion of the total applied,
particularly as it is in the economic interest of the yard
to keep such losses to a minimum. It is likely that losses
from washing/scraping co-polymer paints were also
small. Much of the depleted paint would have degraded
and been washed off during the useful life of the coating
and high pressure hosing was normally sufficient to
remove slime films and the remaining degraded surface
down to sound paint, removing little material that
contained high concentrations of TBT. This would not
have been the case a number of years previously, when it
was necessary to scrape off relatively large amounts of
flaking free-association paints to obtain a sound base for
re-painting. In addition, some shipyards actively took
precautions to prevent contaminated water and chippings from the bottom of the dock from entering
adjacent waterways. It is therefore concluded that the
losses from dry docks would have been small. The
shipping undergoing treatment in dry docks in Europe
each year would use around 48 t TBT, of which it is
estimated (as above) that approximately 1% (0.48 t),
were lost in solution to the environment. By comparison,
Harris et al. (1991) estimated that although a Naval
Dockyard discharged up to 1 t of TBT annually to the
Tamar estuary, it made a relatively small contribution to
the input of TBT in solution, which was dominated by
leaching from hulls.
(b) From dynamic leaching
The wetted area of the world fleet was approximately
60 M m2 (A. Milne, pers. comm.), and vessels typically
spent approximately 250 days sailing (68.5%) and 115
days stationary each year (Stopford, 1988).
If the dynamic leaching rate of TBT from hull antifoulants is taken to have been approximately 4 ìg cm "2
d "1 (Waldock, 1986), approximately 600 t of TBT
would have been leached from vessels under passage to
the world seas. However, the average passage distance
on the high seas is around 4000 miles, whereas a typical
passage distance in the North Sea is only about 400
miles. Therefore, as only 20% of the cargo passed
through the North Sea, and travelled over about 10% of
the route miles, the annual leaching while on passage
into the North Sea may be estimated as 12 t.
(c) From static leaching
Based upon calculations similar to the above, the input
of TBT to the world seas resulting from static leaching
Table 2. Estimated total annual input of TBT to the North Sea.
Source
New building
Dynamic leaching
Static leaching
Dry dock
Ship-breaking
Total
Input of TBT
(t)
0.05
12.00
55.30
0.48
Negligible
67.83
(i.e. 5.6% of the total
world usage of TBT)
while vessels were in port was 276 t. As 20% of world
trade was in the North Sea, 55.3 t of TBT would have
been leached in North Sea ports (this is probably an
upper estimate, since North Sea ports were generally
very efficient and handled cargoes rapidly).
(d) Ship-breaking
Almost no ship-breaking was carried out in northern
Europe (Stopford, 1988), therefore TBT losses through
scrapping will be negligible.
The estimated total input of TBT to the North Sea
from the individual sources is therefore shown in Table
2. Some indication of the reasonableness of these calculations can be obtained by considering that the sum of
the estimated losses of TBT from the world fleet when
static (276 t) and the loss when on passage (600 t) is 73%
of the estimated world annual usage (1200 t) at that
time. These two components are considered to make up
the majority of the input to the environment from
ship-related activities. Some of the shortfall may be
accounted for by the incomplete removal of paint
between repaintings, so that some ‘‘old’’ TBT will accumulate on the ships until scrapping. A relatively small
change in the average leaching rate used in the above
calculations, e.g. an increase from 4 to 5 ìg cm-2 d "1
(Waldock, 1986) could account for almost all the shortfall. It appears, therefore, that the calculations are, in
broad terms, reasonably in balance.
Significance of the estimated inputs to
various areas of the North Sea
The above calculations have been used to derive estimates of the inputs of TBT to the coastal waters of the
North Sea, and also to the wider North Sea from vessels
on passage. The biological impact of these inputs will
have been primarily dependent upon the concentrations
of TBT that were generated in the environment from
these inputs. From Table 3, it appears that the largest
input was to the Netherlands coast. The strongest
*Half time=0.5 years.
1 Northern North Sea
2
3i Northeast UK
3ii Southeast UK
English Channel
4 Netherland/Belgium
7i Central
7ii Central
5 Germany/Denmark
Skagerrak
6 Norway
Sea area
On
passage
6
18
201
373
524
825
21
29
201
79
59
0.60
0.35
2.04
3.91
5.66
26.11
0.00
0.00
7.65
7.49
1.27
0.03
0.08
0.94
1.75
2.46
3.87
0.10
0.14
0.94
0.37
0.28
0.63
0.43
2.98
5.66
8.12
29.98
0.10
0.14
8.59
7.86
1.55
0.50
0.50
0.03
0.01
0.18
0.19
0.60
0.45
0.36
0.20
0.36
Without
With
Shipping
Total
Outflow
movements degradation degradation*
(t)
106 m3 s "1
In
harbour
Local input
(t)
0.75
0.35
0.30
0.30
0.20
0.17
0.35
0.20
0.10
0.10
0.50
0.00
0.08
0.00
1.00
0.00
8.02
29.92
59.40
37.97
30.81
25.66
Flushing From inflow
time
water
(y)
(t)
0.95
0.79
75.59
506.06
34.28
152.00
38.03
100.55
98.29
146.93
57.44
Rank
0.34
0.49
49.87
333.90
25.98
120.09
23.41
76.21
85.57
127.92
28.72
11
10
6
1
8
3
9
5
4
2
7
0.60
0.60
0.03
0.01
0.15
0.16
0.01
0.06
0.18
1.00
1.80
0.40
0.30
3.00
1.30
0.23
0.20
21.00
1.50
0.20
0.10
0.10
0.57
0.66
0.02
0.17
0.73
0.76
0.00
0.13
0.76
0.87
0.87
Exp
0.63
0.43
2.98
5.66
8.12
29.98
0.10
0.14
8.59
7.86
1.55
0.80
0.72
75.59
432.05
41.14
167.21
8.39
74.96
102.87
16.64
6.95
0.46
0.48
1.18
71.29
30.04
126.73
0.00
9.37
77.96
14.49
6.05
10
9
8
3
4
1
11
6
2
5
7
Total
Without
With
(t)
degradation degradation* Rank
Concentration ngl "1
Flushing rates derived from observations
Without
With
Outflow
Flushing
degradation degradation* 106 m3 s "1 time (y)
Concentrations ngl "1
Flushing rates derived from modelling
Table 3. Model of relative concentrations of TBT in various areas of the North Sea, based on flushing rates derived from modelling and from observation (ICES, 1983).
38
I. M. Davies et al.
62°N
1
60°
6
58°
2
3'
56°
7'
7"
5
54°
3"
4
52°
50°
100 m contour
48°
6°W
4°
2°
0°
2°
4°
6°
8°
10°E
Figure 1. Subdivision of the North Sea used in this report, from ICES (1983).
impacts might, therefore, be expected in this area. However, this approach disregards a range of hydrographic
and chemical factors, such as the volume of the receiving
sea water, its rate of advection away from the input area,
and the degradation of TBT in sea water that took place
after release.
Tributyltin inputs to the North Sea
0.6
0.7
39
1.8
1
0.3
0.3
0.03
0.3
2
6
0.6
3'
1.0
7'
Sk
0.8
0.01
0.01
0.18
0.02
7"
0.06
0.01
3"
5
0.04
0.01
0.12
4
0.15
Figure 2. Summary diagram of observed exchange pattern between North Sea areas. Estimated advective fluxes between boxes in
106 m3 s "1 (turbulent exchange indicated by dashed arrows between boxes).
A preliminary approach to assessing the relative TBT
contamination status (and hence risks of biological
effects in sensitive organisms) of areas in the North Sea
has been made using a simple box model of water
circulation in the North Sea, and exchange with the
North Atlantic and Baltic Sea (Fig. 1). In ICES (1983)
the author(s) present a model, which treated the North
Sea as being divided into nine boxes (plus the Skagerrak
and the English Channel), included information on the
flushing times of the boxes, and described the general
40
I. M. Davies et al.
Table 4. Effects of variation of the degradation half-time of
TBT between 2.0 and 0.2 years on the ranking of sea areas.
Half-time
2.0 yr
Area
1 Northern North Sea
2
3i Northeast UK
3ii Southeast UK
English Channel
4 Netherlands/Belgium
7i Central
7ii Central
5 Germany/Denmark
Skagerrak
6 Norway
10
9
6
1
5
2
11
4
3
7
8
1.0 yr
0.5 yr
Rankings
10
9
7
1
5
2
11
4
3
6
8
10
9
8
3
4
1
11
6
2
5
7
0.2 yr
9
8
10
5
3
1
11
7
2
4
6
pattern of water movement produced from both
modelled and observed advective fluxes between the
boxes. (Fig. 2).
This simple model has been used to investigate the
possible relative significance to the areas of the North
Sea of the inputs estimated above. The inputs from
vessels in harbour in each country were assigned to the
appropriate boxes in the model (Table 3). The input to
the Skagerrak arose from trade to Sweden and Finland
and initially appears likely to be an underestimate, as
there was additional trade to countries bordering the
south and east of the Baltic. However, TBT released
from vessels in the Baltic proper, and further north, was
unlikely to enter the Skagerrak or North Sea. The
mean residence time of near-surface water in the Baltic
proper is of the order of 25 yr, and so the great majority
of TBT released in Baltic harbours would degrade
before it was transported to the North Sea. It is therefore likely that the input figures used in the model are
over-estimates of the input from shipping in Baltic
harbours.
The annual input from dynamic leaching (arising
from vessels on passage) was estimated as proportional
to the intensity of shipping movements in the various
areas of the North Sea. Collated annual data on shipping movements are not readily available, and therefore
data were extracted from the Atlas of the Seas Around
the British Isles (MAFF, 1981). These data cover most
of the North Sea and approaches, for a single period in
1976 (checked against periods in 1977–1979). In areas
for which no data were available (the eastern North Sea
north of the German Bight), intensities were assumed to
be proportional to the volumes of trade in adjacent
countries. In this way, it was possible to distribute the
input from ships on passage between the boxes in the
model to reflect the distribution of shipping movements.
For reasons discussed above, inputs to the Baltic from
vessels on passage in the Baltic are unlikely to contribute
to TBT inputs to the North Sea. However, the omission
of any consideration of inputs from vessels on passage
to eastern Baltic ports will result in an under-estimate
of inputs to waters at the entrance to the Baltic
(Skagerrak). This may serve to counterbalance the overestimate arising from the treatment of inputs from
Swedish and Finnish harbours as arising within the
Skagerrak.
It is known that TBT degrades in the marine environment and a degradation half-life of 0.5 yr has initially
been used in the model (Waldock et al., 1990). This
value is of the order of those half-lives observed in
aerobic sediment, rather longer than those in water, and
shorter than those in anaerobic sediment.
In applying this circulation model, the behaviour of
TBT has been described in a very simple manner. TBT is
treated as a degradable dissolved substance, and the
water within each box in the model is considered to be
well mixed. It is assumed that inputs to harbours enter
the main body of the North Sea and therefore can be
considered as having the same potential for environmental impact as direct releases to the open sea from
vessels on passage. It is recognized that TBT interacts
strongly with solids and accumulates in areas of finegrained sediment (Readman and Mantoura, 1990).
Therefore the location of harbours (e.g. on the open
coast, as opposed to in estuarine areas with muddy
sediment) may have considerable influence on the proportion of the TBT from any particular harbour that
reaches the open sea (Harris et al., 1991). However, the
considerable number of harbours in the North Sea area
may serve to reduce any geographical bias in the overall
distribution of inputs arising from details of individual
harbours. Consequently, the model may be able to
indicate the relative degrees of contamination of
large areas, but cannot describe details of distributions of TBT around particular harbours or shipping
lanes.
The concentration of TBT in each box is firstly
calculated as the flux of TBT to the box from static
leaching, and ships on passage, plus that advected in
from neighbouring boxes, divided by the out-going
advective flux of water. This concentration is then
decreased appropriately to allow for the degradation of
TBT during the residence time of water in the box.
The TBT concentrations (Table 3) obtained from
calculations based upon flushing rates derived from
modelling (ICES, 1983) show the least contaminated
areas to have been sub-divisions 1 and 2, in the northern
North Sea, remote from large commercial harbours
(other than Sullom Voe, Shetland). The most contaminated areas are identified as those off the east coast of
England (Area 3ii), and off Belgium, the Netherlands
and Germany (Areas 4 and 5) and the Skagerrak (Area
S). Area 3ii received only a moderate input, but had only
Tributyltin inputs to the North Sea
41
Table 5. Effects on the ranking of sea areas of variation in the assumed efficiency of transfer of TBT
from harbours to the open sea between 100% and 2%, and also of the inclusion of estimated inputs
from dry dock activities for losses of TBT of 0 and 50%.
Transfer efficiency (%)
100
Sea area
50
2
Ranking
1 Northern North Sea
2
3i Northeast UK
3ii Southeast UK
English Channel
4 Netherlands/Belgium
7i Central North Sea
7ii Central North Sea
5 Germany/Denmark
Skagerrak
6 Norway
10
9
8
3
4
1
11
6
2
5
7
10
9
8
2
4
1
11
6
3
5
7
10
9
8
2
4
1
11
6
3
5
7
Table 6. Results of the ranking of the North Sea areas for risk
of biological impact from the model, based on observed flushing times (ICES, 1983), and the measured effects on dogwhelks
(imposex; Harding et al., 1992).
Ranking
3i Northeast UK
3ii Southeast UK
4 Netherlands/Belgium
5 Germany/Denmark
6 Norway
Skagerrak
English Channel
Modelled
risk
Observed
impact
7
3
1
2
6
5
4
7
5
1
3
1
4
6
0
Conc ng l
a small advective flux. The Areas 4 and 5 received much
greater inputs, but also experienced a greater advective
flow, which partially compensated for the input. The
most surprising output from the model is the high
concentrations in Areas 7i and 7ii. The advective flow
out of the former is very small, while the latter receives
a significant input from Area 4 to the south.
If the same calculations are carried out using flushing times from observations (ICES, 1983), a different
pattern emerges (Table 3). The main differences between
the modelled and observed flushing times are that the
observed times are considerably longer in Areas 3i, 3ii,
7i, and 7ii. In this case, relatively low risks of environmental impact are again predicted in Areas 1, 2, and 6,
but now also in Areas 3i, 7i, and 7ii. The greatest risks
are indicated in Areas 4, 5, and 3ii, and intermediate
levels of risk in the Skagerrak and the English Channel.
Area
Losses due to dry docking activities (%)
"1
0.5
0.5
1.2
71.3
30.0
126.7
0.0
9.4
78.0
14.5
6.1
50
Rank
Conc ng l "1
Rank
10
9
8
3
4
1
11
6
2
5
7
0.7
0.7
1.5
92.7
39.1
173.2
0.0
12.7
107.1
20.1
8.4
10
9
8
3
4
1
11
6
2
5
7
Discussion
The estimated concentrations of TBT in the modelled
boxes are all greater than would be expected to be found
for dissolved TBT in North Sea waters (Karbe, 1992)
although Cleary (1991) reported concentrations of TBT
in unfiltered sea water of up to 30 ng tin l "1 in subsurface samples, and 4–75 ng tin l "1 in the surface
microlayer. Quantitative comparisons of this nature may
be misleading. The model does not discriminate dissolved or particulate-borne TBT, nor can it highlight
local elevations of TBT concentrations in water, as have
been found at sources of TBT such as harbours or
marinas (Langstone and Pope, 1995). The results are not
comparable to individual measured values of TBT concentrations in the different parts of the water column.
Such observations are better addressed through other
modelling techniques, which can include interactions of
TBT with particles, and other processes (e.g. Harris
et al., 1991). The absolute concentrations derived from
the current model should not therefore be considered
important, but the results may indicate the relative risk
of biological effects over large areas.
In general, both models (Table 3) highlight areas 3ii
(English coastline) and 4 (Netherlands/Belgium coastline) as those likely to have shown the greatest TBT
concentrations in the water, and therefore with the
highest potential risk to many marine organisms. Area 5
(German/Danish coastline) and the Skagerrak were the
areas with the next highest predicted risk, from both the
observed and modelled flushing rates.
To explore the robustness of the results of the
modelling, the sensitivity of the output to the values of
some of the important parameters of the model was
investigated.
42
I. M. Davies et al.
Firstly, a degradation half-life of 0.5 yr for TBT
(Waldock et al., 1990) was used in the calculations. TBT
has an affinity for particulate matter, resulting in
sediments being the major reservoir for TBT (Langstone
and Pope, 1995). Very high degradation rates of dissolved TBT (7–15 d, Seligman et al., 1986) were considered inappropriate, and therefore a lowest half-life
of 0.2 yr was adopted. The effects on the relative risk
of biological effects of varying the half-life of TBT
between 0.2 and 2 yr are shown in Table 4. While some
areas do change in relative risk, the general pattern of
risk is not greatly dependent upon the assumed half-life
of TBT.
Secondly, it was assumed that TBT released in
harbour areas was as available to the open sea as TBT
released directly from vessels on passage. As discussed
above, the affinity of TBT for particles will cause a
proportion of the TBT to have been retained within
harbours and estuaries. The effect on the ranking of sea
areas of varying the assumed efficiency of transfer of
TBT from harbours to more open sea areas between
100% (as in the initial model results, Table 3) and 2%, is
shown in Table 5. It can be concluded that, although the
assumed efficiency of transfer affects the estimated TBT
concentrations, the ranking of the individual sea areas is
not affected.
Thirdly, it has been assumed that the leaching rate
of TBT from stationary vessels is equal to that from
vessels on passage. Dynamic leaching tests showed
5–10 ìg cm "2 d "1 leaching rate for copolymer paints
with a proposed manufacturers leach rate of approximately 1 ìg cm "2 d "1 (Waldock, 1986). It is likely
that leaching rates are greater when vessels are
underway, although direct measurement is difficult
(Schatzberg, 1990). Leaching rates have been estimated
from experiments to range from 0.11–6.75 ìg TBT
cm "2 d "1 depending on the type of paint (conventional or copolymer) and the length of immersion in
aerated sea water (Anderson and Dalley, 1986). Other
studies (Waldock, 1986) using a constant flow of
water over a 2–3 mo period showed leaching rates of
4–5 ìg cm "2 d "1, illustrating the variability of leaching rates in the literature. Indeed, Harris et al. (1991)
used a leaching rate of 0.1 and 0.3 ìg cm "2 d "1 when
modelling the Tamar estuary, and acknowledged that
this may be as much as an order of magnitude too
low. Literature values for leaching rates are therefore rather variable. If the leaching rate is lower in
harbours, this can be considered equivalent, so far as
the model is concerned, to a reduction in the apparent
transfer efficiency of TBT from harbours to the
open sea. As indicated above, variation over a
wide range of values of transfer efficiency (100–2%),
equivalent to varying static leaching rate from
4–0.1 ìg cm "2 d "1, has little effect on the conclusions
from the model.
Finally, the losses to the environment from drydocking were considered to be insignificant (0.48 t), and were
omitted from the modelling. It may be that this is an
under-estimation of the actual losses, and the sensitivity
of the modelling results to this parameter was therefore
investigated. Dry docks are generally situated in port
areas and the distribution of inputs of TBT from dry
docks may therefore be modelled in the same way as the
inputs from vessels in port, i.e. in proportion to the
distribution of seaborne imports between countries. It is
then possible (Table 5) to calculate the effect of varying
the assumed percentage loss of TBT from dry docks
from 0% (as in the initial model) to 50%. Losses of 50%
increase the calculated concentrations of TBT, but have
no effect on the ranking of the sea areas. It appears
therefore that uncertainty in the scale of input of TBT
from dry docks has no effect on the conclusions of the
model.
It may be concluded that the main simplifying
assumptions concerning the inputs and behaviour of
TBT do not greatly affect the general conclusions from
the model.
The relative risk of biological effects of TBT predicted
by the model have been compared with the results of a
broad study of the biological effects of TBT around the
North Sea (Harding et al., 1992). These authors studied
imposex (the development of male sexual characteristics
in female dogwhelks, Nucella lapillus, in response to
TBT exposure, Bryan et al., 1986) around the coastline
of the North Sea. The Vas Deferens Sequence Index
(VDSI, the stage of development of the male sexual
organs, Gibbs et al., 1987) was measured. The results
allowed each population examined to be classified as
unlikely to show any reduction in reproductive capacity
(VDSI <4); or as populations with higher impact of
TBT, a proportion of sterile females and reduced reproductive capacity (VDSI 4–6). For the purposes of comparison, the dogwhelk sampling sites were grouped using
the same geographical boxes as in the model. The
percentage of dogwhelk sampling sites in each area at
which the dogwhelks showed VDSI values of 4–6
was calculated, and the percentages ranked and compared to the results of the input model (Table 6).
The predominantly offshore areas 1, 2, 7i and 7ii
were not sampled for dogwhelks, and are therefore
excluded.
In broad terms, the actual measured biological effect
mirrors the predicted areas of highest risk biological
effect from TBT from derived shipping. There are two
areas where the results differ, area 3(ii) along the
English coast, and area 6 the western Norwegian coast.
In area 6, the majority of the sites were chosen to
investigate the extent of TBT contamination in
harbours, 8 out of the 12 sites sampled in this area
were from busy harbours. In this way, the sampling
may have been biased and over-estimated the general
Tributyltin inputs to the North Sea
extent of TBT contamination along the coastline. The
imposex results in area 3(ii) were based on relatively
few samples.
Conclusions
1. An assessment is presented of the distribution of
inputs of TBT to the North Sea from shipping
activities.
2. Combination with a simple box model of water
circulation in the North Sea allows prediction of
relative risk of biological effects caused by TBT in
different areas of the North Sea.
3. In broad terms, the predicted relative risks reflect the
observed distribution of effects in Nucella lapillus.
Acknowledgements
The data used in the calculations of inputs were compiled by Mr A. Milne, formerly of Courtaulds NCT, and
Dr W. Turrell suggested the use of the hydrographic
model, to whom the authors extend their sincere thanks.
References
Anderson, C. D. and Dalley, D. 1986. Use of organotins in
antifouling paints. Proceedings of the First International
Organotin Symposium, Washington DC, 23–25 September
1986, pp. 1108–1116.
Bailey, S. K. and Davies, I. M. 1988. Tributyltin contamination
around an oil terminal at Sullom Voe, Shetland. Environmental Pollution, 55: 161–172.
Bryan, G. W., Gibbs, P. E., Hummerstone, L. G., and Burt,
G. R. 1986. The decline of the gastropod Nucella lapillus
around south west England: Evidence for the effect of
tributyltin from antifouling paints. Journal of the Marine
Biology Association of the UK, 66: 611–640.
Cleary, J. J. 1991. Organotin in the marine surface microlayer
and subsurface waters of south-west England: relation
to toxicity thresholds and the UK Environmental Quality
Standard. Marine Environmental Research, 32: 213–222.
Gibbs, P. E., Bryan, G. W., Pascoe, P. L., and Burt, G. R. 1987.
The use of the dogwhelk Nucella lapillus, as an indicator of
43
tributyltin (TBT) contamination. Journal of the Marine
Biology Association of the UK, 67: 507–523.
Harding, M. J. C., Bailey, S. K., and Davies, I. M. 1992. TBT
imposex survey of the North Sea. Scottish Fisheries Working
Paper, 9/92. 26 pp.
Harris, J. R. W., Hamlin, C. C., and Stebbing, A. R. D. 1991.
A simulation study of the effectiveness of legislation and
improved dockyard practice in reducing TBT concentrations
in the Tamar estuary. Marine Environmental Research, 32:
279–292.
ICES. 1983. Flushing times of the North Sea. International
Council for the Exploration of the Sea. Cooperative
Research Report No. 123.
Karbe, L. 1992. Toxicity of surface microlayer, subsurface
water and sediment-elutriates from the German Bight:
summary and conclusions. Marine Ecology Progress Series,
91: 197–201.
Langstone, W. J. and Pope, N. D. 1995. Determination of TBT
absorption and desorption in estuarine sediments. Trace
metals in the aquatic environment: Proceedings of the 3rd
International Conference, Aarhus, 16–20 May, 1994. Phillips
D. J. H. and Rainbow P. S. (eds).
MAFF. 1981. Atlas of the sea around the British Isles. Ministry
of Agriculture, Fisheries and Food.
Readman, J. W. and Mantoura, R. F. C. 1990. Partitioning of
butyltins in coastal waters and sediments. 3rd International
Organotin Symposium, Monaco, 17–20 April 1990.
Schatzberg, P. 1990. Tributyltin release rate measurement and
significance. 3rd International Organotin Symposium,
Monaco, 17–20 April 1990.
Seligman, P. F., Valkirs, A. O., and Lee, R. F. Degradation of
tributyltin in marine and estuarine waters. Proceedings of
the Conference Oceans 86, Volume 4, Organotin Symposium,
Washington DC, 23–25 September 1986.
Stopford, M. 1988. Maritime Economics. Unwin Hyman,
London.
Ten Hallers-Tjabbes, C. C., Kemp, J. F., and Boon, J. P. 1994.
Imposex in whelks (Buccinum undatum) from the open North
Sea: relation to shipping traffic intensities. Marine Pollution
Bulletin, 28: 311–313.
Waldock, M. J. 1986. TBT in UK estuaries, 1982–96. Evaluation of the environmental problem. Proceedings of the
Conference Oceans 86 volume 4, Organotin Symposium,
Washington DC, 23–25 September 1986.
Waldock, M. J., Thain, J. E., Smith, D., and Milton, S.
1990. The degradation of TBT in estuarine sediments. 3rd
International Organotin Symposium, Monaco, 17–20 April
1990.