Prepared for Shell U.K. Exploration and Production by Rudall Blanchard Associates Limited
9. CONCLUSIONS AND SUMMARY IMPACT HYPOTHESIS
REFERENCES 30
1.1. The Brent Spar was installed in the North Sea Brent Field in 1976, to provide a storage and tanker off-loading system for the Brent platforms: The buoy is approximately 140 m in height, cylindrical in shape and moored to the sea bed by six anchors. it is made up of large oil storage tanks at the base, buoyancy tanks towards the middle and topsides containing the offshore tanker loading equipment.
1.2. The Spar reached the end of its useful working life in the late 80's and was taken out of commission in 1991, in preparation for abandonment. Various onshore and offshore abandonment options have been examined by the operators, Shell (UK) Exploration and Production. The preferred option is to tow the structure to an approved deep water disposal site in the NE Atlantic and sink it. A summary of the selection process can be found in the Best Practicable Environmental Option Assessment Report for the Brent Spar abandonment.
1.3. This report describes the predicted fate of the Spar based on the characteristics and contents of the buoy, and current knowledge of the deep sea environment. The term used to describe this type of prediction in the '1991 Oslo Commission Guidelines for the Disposal of Offshore Installations at Sea' is an 'Impact Hypothesis'.
1.4. The environmental impacts of three sinking scenarios have been considered. The most likely scenario is that the buoy will sink to the sea bed and remain intact after impact. This means that the residual contaminants contained within the structure will slowly leak into the bottom waters arid sediments ('slow release' scenario). The alternative of major damage to the storage tanks (and fast release of residual contaminants contained within the structure 'fast release' scenario) on impact is considered less likely to occur. The third alternative of tank damage (and release of some of the residual contaminants in the near surface environment - 'surface release' scenario) is considered unlikely to occur. In all cases, corrosion of the structure will slowly release metals into the water column.
1.5. The potential environmental impacts of each sinking scenario are discussed and the Impact Hypothesis for the most likely scenario (i.e. 'slow release') is developed. For all three scenarios the hull of the buoy will remain intact for several thousand years due to the slow rates of corrosion. Eventually the structure will break up and become buried in the sediment system. The potential environmental impacts from the physical presence of the structure will be highly localised.
1.6. In the 'slow release' scenario, elevated levels of hydrocarbon and heavy metal would be concentrated in the sediment close to the structure, although hydrocarbon may be detectable up to c.500 m and biological effects up to 150 - 250 m from the structure. The effects of the 'rapid release' scenario, would follow a similar pattern although the area affected would be expected to be larger (hydrocarbon may be detectable up to c.l000 m and biological effects up to 500 m from the structure) but the effects would be over a much shorter duration. Given the small numbers of organisms present in the deep ocean and low encounter rates with the affected area, effects will be insignificant. In the third and least likely scenario (i.e. 'surface release'), the overall effects on the benthos would be similar to those in the two scenarios. The additional impacts of the surface release on pelagic organisms in the 'sinking zone' would be insignificant.
1.7. Conservative modelling of the long-term fate of the low specific activity scale. indicates that the estimated dose rates to individuals (human) would be several orders of magnitude less than those regarded by the IAEA as de minimus for the purposes of regulatory control. Similarly, dose rates to human populations would be well within those acceptable to IAEA. The dose rates to deep sea organisms would be within the ranges encountered in nature and effects on natural populations would not be expected.
2.1. The selection of deep sea disposal as the preferred abandonment option for the Brent Spar as been described in the Best Practicable Environmental Option Assessment Report. The present document develops and describes the Impact Hypothesis for the disposal of the buoy at an authorised site in the North East Atlantic.
2.2. The Brent Spar, and a summary of the inventory of materials associated with the structure, is described in Section 3. A synopsis of the proposed abandonment operation is provided in Section 4 and the main features of the potential disposal sites in the NE Atlantic outlined in Section 5.
2.3. The fate of the structure from the time it is sunk to the time it eventually breaks up and becomes incorporated into the sediment system is discussed in Section 6. Three possible scenarios for the fate of the structure during the initial sinking are considered.
2.4. The fate and effects of the residual hydrocarbon and heavy metal contaminants contained within the structure are described in Section 7 and the long-term fate and effects of the small amounts of Low Specific Activity Scale (LSA) considered in Section 8. More details on the mathematical models used in the long term predictions of LSA scale are included in an Annex.
2.5. The conclusions are presented in the form of a short Impact Hypothesis in Section 9.
3.1. DESCRIPTION OF THE FACILITY
3.1.1. The Brent Field is located in Block 211/29 of the UK Sector of the North Sea, some 190 km ENE of Shetland. The Spar is located at Latitude 61degų03' 14.7" N, Longitude 01degt40'04" E, approximately 2.9 km from the Brent 'A' platform and 1.9 km from Brent 'B' platform the water depth in the field is some 140m. [Figure omitted]
3.1.2. The Brent Spar is a 29m diameter cylindrical buoy which floats vertically in the water with a draft of 109m and a height above water of 28m. It is moored to the sea bed by six anchors. It was constructed using techniques similar to those of ship building and the main structure consists of a thin outer skin of 20 mm plate steel, stiffened by ribs and bulkheads.
3.1.3. The upper section consists of a helideck, crane, tanker mooring and loading boom, and accommodation for 30 people. In the middle of the structure are twelve buoyancy tanks. Below this, the main storage area is divided into six tanks which extend to the base of the buoy, and have a total capacity for c. 300,000 barrels of oil. At the base of the buoy, there is a sealed compartment containing the permanent ballast.
3.1.4. The facility was used to store oil produced at the Brent 'A' platform and to act as a tanker loading facility for the Brent Oil Field. Installed in 1976, it was the sole route for the export of crude oil until the Brent System Pipeline was commissioned in 1978. Since 1978, the Spar has continued to be used as an alternative export route.
3.1.5. The costs of maintaining the Spar increased substantially in the period 1987 to 1990. In 1991, a review concluded that the work necessary to refurbish the facility. and extend its operational life. would cost up to £90 million. The buoy would have to be taken out of commission for a 2 - 3 year period during the refurbishment. Given the age of the structure, the presence of a pipeline system for the export of crude oil and the substantial cost; of refurbishment, it was decided that the Brent Spar should cease operations. It was taken out of commission in September 1991 after l5 years service, in preparation for abandonment. [Figure omitted] Page 5 Brent Spar Abandonment - Impact Hypothesis - 15/12/94
3.2. INVENTORY OF MATERIALS
3.2.1. The total weight of the structure is 14,500t, made up of 6,700t of structural steel. 6,800t of permanent ballast and 1000t of equipment. The permanent ballast is solid and made up of haematite, embedded in concrete. Like steel, this material is considered to be inert.
3.2.2. Of particular relevance to the eventual disposal of the structure are the materials (e.g. heavy metals. oils etc.) which might pose a threat to the environment. Inventory items have been classified in accordance with the "Convention for the Protection of the Marine Environment of the North East Atlantic 1992" (Oslo and Paris Commissions - OSPARREV). This Convention regulates substances and materials whose disposal at sea is controlled and are referred to as "regulated substances". Surveys carried out in 1991 identified a number of these substances aboard the Spar (AURIS, 1994; Metocean, 1993). These are dominated by three generic classes of 'regulated substances', notably the petrogenic hydrocarbons, heavy metals and low specific activity scale, although there are small quantities of other classes on the structure.
3.2.3. Equipment throughout the structure contains small quantities of heavy metals (in metallic form) as an integral part of the materials of construction. The sacrificial anodes contain an estimated 28.7t of aluminium, 10.2t of zinc metal and minor quantities of cadmium (8 kg), lead (0.6kg), and mercury (0.lkg). Electric cables contain some 13.5t of copper and paint on the structure includes some 3.5t of zinc. Small quantities of lead (2.5kg) and of nickel (3.5kg) are contained in remaining batteries.
3.2.4. Small quantities of other materials are also contained in some of the equipment. Although PCB-containing transformer fluids were replaced some years ago, traces of PCB (<20ml) may still remain in the two transformers. Synthetic materials and plastics are also located on the structure (e.g. in fittings in the accommodation/control room and cable insulation) but the exact quantities are not known (AURIS, 1994; Metocean, 1993).
3.2.5. The oil storage tanks contain some 48,000m3 of sea water together with an estimated 100t of oily sludge at the bottom. A 1991 analysis indicated that the oily sludge contains an estimated 9.2t of oil and a number of heavy metals, including cadmium (5.8kg), chromium (2.1 kg), copper (42.9kg), nickel (3.9kg), lead (8.9kg), zinc (87.4kg), arsenic (0.3kg), and mercury (0.2kg). The remainder of the sludge is composed of a mixture of sand and scale. The walls of the storage tanks are also coated with an estimated 41.3t of hydrocarbons, in the form of thin layer of oil and wax.
4.2.6 The sea water in the storage tanks has not been analysed in detail and it has been conservatively assumed that it will contain hydrocarbons (up to 40ppm) from the residual oil in the tanks together with zinc (up to 12ppm) and aluminium (up to 19ppm) from the internal anodes (AURIS, 1994). In 1991, the storage tanks were treated with 4,500 litres of glyoxal to remove hydrogen sulphide which had accumulated within them. Glyoxal reacts with hydrogen sulphide to form various thiols. It is believed that all the glyoxal reacted in this way and that all the reaction products were removed from the tank during the 1991 decommissioning operation (AURIS, 1994). Page 6 Brent Spar Abandonment - Impact Hypothesis 15/12/94
3.2.7. The location and estimated quantities of the main organic and heavy metal contaminants within the structure are summarised in the table. Contaminant Amount (kg) Sludge Anodes Paint BatteriesCables Lights Tank Total & Walls Electrics ORGANICS PCBs - - - Trace - - - Trace Hydrocarbon 9,400.0 - - - - - 41,300.0 50,700.0 METALS Aluminium - 28,677.0- - - - - 28,677.0 Arsenic 0.3 - - - - - - 0.3 Bismuth - 29.0 - - - - - 29.0 Cadmium 5.8 8.0 - 2.6 - - - 16.4 Chromium 2.1 - - - - - - 2.1 Copper 42.9 - - - 13,500.0 - - 13,542.9 Indium - 10.2 - - - - - 10.2 Lead 8.9 0.6 - - - - - 9.5 Mercury 0.2 0.1 - - - Trace - 0.3 Nickel 3.9 - - 3.5 - - - 7.4 Silicon - 48.0 - - - - - 48.0 Titanium - 8.8 - - - - - 8.8 Zinc 87.4 10,224.0 3,500.0 - - - - 13,811.4
3.2.8. Scale is commonly found in oil processing facilities throughout the world. In many areas scale may be contaminated by deposition of small amounts of naturally occurring radioactive salts from the oil reservoirs formation to form low specific activity (LSA) scale. The sludge and scale in the Brent Spar contain a number of naturally occurring radio nuclides, some of which have very short radioactive half-lives (seconds, minutes or hours) and others which are longer-lived (half-lives of years). These radio nuclides are all members of "chains", that is they are formed by the decay of radioactive parents and themselves decay into radioactive daughters (except for the last chain member which is stable). In order to estimate the radiological impact of sea disposal of the installation, it is necessary to know the quantities of radio nuclides present. particularly those with longer half-lives. The relevant radio nuclides and their half-lives are: radium-226, 1600 years; lead-210, 22.3 years; polonium-210, 138 days; radium-228, 5.75 years; thorium-228, 1.91 years.
3.2.9. Measurements have been made of the concentrations of some radio nuclides in the sludge in the Brent Spar storage tanks. In the case of scale, radio nuclide concentrations can be derived from measurements made on material from the Brent Alpha and the Brent Bravo. The average activity of radium-226 and actinium-228 in the sludge is 4.5 and 3 Bq g-1. respectively. The average activity of radium-226 and actinium-228 in the hard scale is 17.6 and 15.2 Bq g-l, respectively. in the pipework. Conservatively. taking the total masS OFsludge on the Spar to be 100 tonnes and that of scale to be 30 tonnes, and assuming that all the radio nuclides present are in secular equilibrium, gives the following results for the quantities of the longer-lived radio nuclides in both sludge and scale (see Annex for further details).
3.2.10. These quantities can be placed in perspective by comparing them to calculations by the International Atomic Energy Agency (IAEA) as part of its work to provide the categories of radioactive waste for the purposes of the London Dumping Convention (IAEA, 1986). The IAEA calculated the quantities of radio nuclides which could be disposed of in the deep ocean each year for 1000 years without giving rise to radiation doses exceeding the limit recommended by the International Commission on Radiological Protection (ICRP). The quantities in the Brent Spar sludge and scale are less than one millionth of those calculated by the IAEA.
3.2.11. For further perspective, a comparison can be made with levels of radio nuclides which are naturally present in sea water. Using ranges of sea water concentrations of radium-226, lead-210, polonium-210, radium-228 and thorium-228 (Pentreath, 1988; IAEA, 1988a), it can be shown that the total quantities of these radio nuclides in Brent Spar sludge and scale are of the order of one hundred millionth to one millionth of those in the water of the North East Atlantic. Page 8 Brent Spar Abandonment - Impact Hypothesis 15/12/94
4.1. Thirteen possible methods of abandonment or re-use were initially put forward for consideration, of which six were identified as viable options:
4.2. Of these, horizontal dismantling and deep water disposal were considered in detail to determine the Best Practicable Environmental Option (BPEO) for disposal. The BPEO Assessment demonstrated that the preferable option would be to dispose of the Brent Spar at an authorised deep water disposal site, on the grounds of reduced technical risk; the reduced safety risk to the workforce; the insignificant environmental impact; and the total cost.
4.3. Prior to deep sea disposal the Spar would be 're-entered', the topsides made safe and the state of the structure evaluated. To minimise the potential environmental impacts at the deep water disposal site, a clean-up operation involving trained personnel would remove accessible hazardous materials for disposal onshore. This would include fuel oil, lubricants, household items. and where possible, heavy metals and synthetics.
4.4. Explosives experts would then board the buoy and place the charges around the ballast tanks, which would eventually be used to sink Brent Spar. The number, weight, type and position of the charges installed would be sufficient to ensure that all air filled compartments flooded, even if some charges failed to detonate.
4.5. Two tugs accompanied by an attendant survey vessel would then tow the buoy to the selected deep sea disposal site, using two of the anchor chains as tow lines. The remaining chains would be used as emergency tow lines. Three general areas have been identified by the Scottish Office Agriculture and Fisheries Department (SOAFD) as potentially suitable for the disposal of redundant offshore structures. These are located within U.K. waters and lie in water depths in excess of 2000 m. [Figure omitted]
4.6. The tow would be conducted using conventional maritime practices. All routes from the Brent Field to the disposal sites pass to the north of the Shetland Islands and to the north west of the Outer Hebrides. Whichever site is selected, the earlier stages of the planned routes would essentially be common. The first part of the tow, to the 200 m contour, would be subject to detailed survey. to ensure that the buoy did not encounter any unexpected obstructions. The route for the latter stage of the tow would depend on the dump site selected: The precise routing would be determined by the requirements for adequate water depth, sufficient clearance from installations and the need for sufficient sea room for manoeuvre should bad weather or accident result in breaking of the-tow.
4.7. Prior to sinking the buoy the exact and correct position would be confirmed. The tow vessels would then mechanically release the tow lines and stand off for the detonation of the explosives by remote control. The charges would be detonated simultaneously, to open the buoyancy tanks to the sea. This would result in rapid flooding of all the empty tanks which would remove the remaining buoyancy and sink the structure.
5.1. DISPOSAL SITE LOCATION
5.1.1. The location of the three potential deep water disposal sites, identified by SOAFD, are:
5.1.2. A detailed survey was carried out in August - September 1994 by Shell Expro and SOAFD, to determine the suitability of the proposed disposal sites and, if possible, to identify the optimum site for disposal of the Brent Spar. The survey also collected baseline data against which any future consequences of the disposal operations could be monitored.
5.2. PHYSICAL ENVIRONMENT
5.2.1. The water in the North East Atlantic is continually moving in response to wind, temperature, and salinity gradients. The flow of the currents set up are moderated and directed by the physical presence of land masses, the contours of the sea bed and other bodies of water. The scale of movement is massive and the water moves with immense inertia. The inflow of water into an area is always counterbalanced by a corresponding outflow and at one position, water at different depths may be moving in completely different directions.
5.2.2. The circulation is influenced by the north easterly movement of warm, saline North Atlantic Drift on the surface, into the Norwegian Sea, and the outflow of dense, cold, water at depth from the Norwegian Basin to the south and west into the Iceland Basin and Rockall Trough.
5.2.3. The continental shelf in the NE Atlantic slopes fairly gently down to a depth of c. 200m. At the seaward edge of this shelf, the downward gradient increases on the continental slope. The base of the slope gives way to the continental rise which leads down to the abyssal plain, c. 4km deep. The plain may be relatively featureless, with a slope of 1:1000. In places it may be interrupted by sea mounts (extinct volcanoes), that do not reach the surface of the water. These features correspond to the ecological zones illustrated. Page 11 Brent Spar Abandonment - Impact Hypothesis 15/12/94 [Figure omitted]
5.2.4. The surface waters are isolated from the deep water by the permanent thermocline below which the temperature falls relatively slowly with depth to below 4degC. The depth of the thermocline is generally in the region of 1000m. Above this zone, the water is well mixed and relatively homogeneous. Near the sea bed, there is a layer of homogeneous water, bounded at its upper limit by the Benthic Boundary Layer. This is formed due to the friction of the sea bed against the water and its thickness depends upon the topography of the sea bed but it is generally restricted from 10 - 100 metres above the sea bed (Richards, 1990).
5.2.5. There is relatively sparse information on the bottom current flow in the NE Atlantic. Current velocities are generally lower than those found in surface waters and are typically < 10 cm/sec. For example. in the Rockall Trough, the mean current velocities were only 1 - 2 cm/sec (Metocean, 1993). The flow of cold water from the Norwegian Basin is not constant and is augmented by the semi-diurnal tidal influence, which can increase the current velocity up to 20 cm/sec. The currents are also subject to periodic reversals as the result of persistent mesoscale eddies, thrown off by strong surface flows, like the Gulf Stream. These are large scale phenomena (c. 50-200km) and may persist for two years or so. These eddies produce current velocities of up to 50 cm/sec, giving rise to 'benthic storms' in which vast quantities of sediment may be picked up and re-deposited (Hollister and McCave. 1984).
5.2.6. The rates of sedimentation and the composition of the sediment depends upon the depth of the water. the proximity to land. the productivity of the overlying ocean and the latitude. Siliceous oozes are ubiquitous in the deeper parts of the ocean and are formed from deposition of diatoms. Radiolarian oozes are found in the tropics. and calciferous oozes are formed from foramaniferans in (relatively) shallow water. Within non-productive water masses (e.g. oligotrophic gyres), the main contributors to sedimentation are volcanic and wind blown desert dust which form the so-called 'red clays'. The rate of sedimentation varies between 0.1-0.2 cm per thousand years for the red clays. to 20cm per thousand years on the continental shelf .The mean sediment depth in the North Atlantic is 300-600m, with 1200m in the basins.
5.2.7. Sediment can move downslope due to sediment slides and slumps. debris flows and turbidity currents. Slides and slumps may occur on slopes as low as 2 degrees and may involve large volumes of sediment. Debris flows, the slow movement of sediment downslope, occur on slopes of less than 0.5 degrees. Turbidity currents are high velocity density currents that carry sediment and water down the continental rise and out onto the abyssal plain. There Iis evidence that these phenomena account for the rapid movement of huge amounts of sediment over long distances (Nardin et al., 1979).
5.3. LIVING RESOURCES
5.3.1. Primary production takes place in the upper water of the ocean. In the North Atlantic the diatoms are the main primary producers and their production peaks twice a year, in the early and late summer (Mauchline, 1986). The small plants are grazed by zooplankton, composed predominantly of crustaceans (e.g. copepods and krill), whose numbers peak in the summer, slightly later than the primary producers. These in turn are preyed upon by carnivorous zooplankton, such as medusae, ctenophores, chaetognaths, gymnosomatous and pteropods. Both the herbivorous and carnivorous zooplankton provide the major food source for fish.
5.3.2. The abundance of fish depends upon the quantity of food available to them. In general, the deep ocean is not as productive as coastal and shallow waters of the Continental Shelf. Commercially exploited fish in the NE Atlantic include blue fin tuna, redfish, poutassou and various gadoids (ICES, 1992).
5.3.3. At any deep water site, the 'ecosystem' overlying it may be divided into zones differentiated principally by depth. The deeper water fish, living below about 1500 m, are relatively isolated from those above and there is little vertical migration across this boundary. The numbers and density of the surface fish will be determined largely by the production in the near surface water, The deep water animals are, with certain exceptions, entirely dependent, both directly and indirectly, upon the downward flow of material derived from the surface water above (e.g. corpses, faecal pellets etc.).
5.3.4. The fish species found at depth are different from those in shallower water and are adapted to life in the high pressure, dark deep sea environment. The biomass found in oceanic waters decreases with depth so that at 4 km there is only in the order of 1% of that found near the surface. However, within a 100 m or so of the ocean floor, the biomass of animals increases again (Wischner, 1980).
5.3.5. The animals on or near the bottom may be mobile, fixed to the bottom or live in the sediment and many different groups are represented. The animals found may be artificially grouped by decreasing size into the megafauna (also known as epifauna), macrofauna, meiofauna and microfauna. The most conspicuous mobile animals of the megafauna are the fish which are dominated by-the macrourids (rat tails), then the brotulids, sickleheads, deep sea cod and benthopelagic eels. All tend to be general feeders, consuming the epifauna but not the infauna (Gordon, 1987). Also included in this group are the Porifera (e.g. glass sponges), Ceolenterata (e.g. sea firs, sea pens, sea fans and the soft, black and horny corals), Echinodermata (e.g. sea lilies and feather stars), Crustacea (e.g. giant amphipods and decapod crustaceans), together with various species of Annelida, Bryozoa, Brachiopodia and Ascidiacea.
5.3.6. The macrofauna are dominated by the polychaetes (bristle worms), together with the crustaceans and molluscs. The meiofauna are made up of both multi-celled and single celled animals. Of the former, the Nematoda dominate followed by small benthic copepods and ostracods. Of the latter, the most diverse are the Foraminifera (of which there are 6000 known species) and the xenophytophores (giant protozoans). The smallest of all the animals are found in the most superficial layers of sediment and comprise of various flagellates. sporozoans, ciliated amoeba and yeast cells.
5.3.7. The numbers of birds found over deep water are markedly lower that those found nearer shore. The species expected would be Leach's petrel, storm petrel, fulmar, kittiwake and puffin. The highest numbers of birds are found over the shelf break, on the margin between deep water and the edge of the continental shelf, and are generally found during the winter and the early spring. Page 13 Brent Spar Abandonment - Impact Hypothesis 15/12/94
5.3.8. Although there are records of seals from far offshore areas, such as Rockall, the only mammals generally found far offshore are whales and dolphins. These animals tend to follow their food and undertake seasonal migrations to the rich feeding grounds in the Arctic and Antarctic waters and to specific breeding grounds in less extreme latitudes. Although widespread in the North Atlantic, they do frequent specific migratory routes. The most used one runs north/south offshore from the continental shelf (Metocean, 1993). The mammals that will use this are the Blue, Fin, North Atlantic Right, Sei, Sperm, Bottle-nosed, Pilot, Humpback and Killer whales.
5.4. RESOURCE USE
5.4.1. One of the main considerations during the selection of potential disposal sites was the use made of the area by other legitimate users of the sea. Potential resource uses in the area of the disposal sites are restricted to fishing, shipping, oil and gas exploration and production, deep sea mining, military activities and sub-sea cables (Metoceon, 1993).
5.4.2. Most fishing activities in these water depths are confined to the upper c. 1000m, although equipment to extend this to c. 1500m is being developed (SOAFD - Personal Communication): Also, most frequently used shipping routes across the Atlantic do pass over the proposed disposal site areas.
5.4.3. Oil and gas exploration and production activities are currently confined to the shallower continental shelf regions. As resources in these areas are depleted, activities will inevitably extend to prospects in deeper water. Sedimentary basins which might contain oil and gas underlie the two inner disposal sites. However, the area of the disposal sites will be small in comparison to the area of these basins and the potential for future conflict small.
5.4.4. The only other mineral resource that has been seriously considered for exploitation is deep sea mining for manganese nodules. The distribution of manganese nodules is highly variable but the most extensive fields lie in the Pacific, and the potential for the North Atlantic appears to be very low. As yet no commercial mining operations have been started anywhere.
5.4.5. No military exclusion zones are located in the disposal site areas. However, the Rockall disposal site is within an area used for submarine exercises.
5.4.6. There are a number of existing and proposed sub-sea cables that cross the disposal site region. However, the location of the disposal sites have been selected to avoid these. Page 14 Brent Spar Abandonment - Impact Hypothesis 15/12/94
6.1. SINKING SCENARIOS
6.1.1. The disposal operation has been designed to ensure that the buoy sinks in the vertical, without any tanks imploding or any debris breaking away. Engineering studies indicate that the structure should remain largely intact upon impact with sea bed, although some damage may occur at the base. The main effects of this would be on the permanent ballast compartment and possibly the lower part of the storage tanks. This would have no affect on the solid ballast (made up of haematite embedded in concrete) but could initiate the slow release of the contents of the storage tanks, immediately after impact. While this is the most likely scenario, there is always some uncertainty in an operation of this type.
6.1.2. For the purposes of the Impact Hypothesis, three possible scenarios are considered as illustrated. .
NOTE 1: Should some of the air filled compartments not be flooded as planned, a local rupture could occur during descent. This would affect the air containing compartment but not the main storage tanks which should remain filled with sea water, throughout. This scenario relates to a hypothetical situation of partial release of the storage tank contents into surface waters. A worst case assumption of a 10% loss of the total tank contents (i.e. equates to over a 50% loss from one tank) has been taken for demonstrative purposes .
6.2. IMPACT OF THE STRUCTURE ON THE SEA BED
6.2.1. The buoy is likely to hit and penetrate the sea bed with the base and then topple on to its side. As the structure approaches the sea bed, the surge preceding it and the impact itself will displace sediments from the impact zone. The sediment cloud produced would drift with the near bottom current while the particles gradually settled back onto the sea bed again. The area affected by the initial impact and subsequent re-deposition of sediment is difficult to forecast accurately.
6.2.2. The minimum area affected by impact would be equivalent to the area of the base of the buoy (approximately 700 m2). plus the area of the buoy when lying on its side (probably up to a maximum of 2,800 m2). Assuming that only the top 10 cm of sediment is displaced to the water column, then some 350 ml of sediment may be re-suspended during the initial impact. Given that most of this material is mud or clay, with settling velocities in the region of 0.002 m/s. current velocities of 0.1 m/s and a sediment cloud of up to 10 -20 m high, most of the material would be expected to resettle within 500 - 1000 m of the point of impact. Studies on disposal of dredged materials confirm that sea bed sediment clouds produced from dumping activities generally persist for short periods of time (Pequegnat, 1983).
6.2.3. Sea bed animals living in the impact zone would be damaged or killed by physical burial or gross disruption of their habitat. Those immediately adjacent to the impact zone (i.e. within 1 -200 metres or so) may be smothered by the re-sedimenting particles. Outside this area, as the cloud spreads radially outwards and becomes more dispersed, any effects on benthic communities will decrease and are unlikely to be detectable beyond one to two hundred metres from the point of bottom contact.
6.2.4. After the impact and resettlement of the sediment, re-colonisation of the newly deposited sediments and the new hard surfaces of the Brent Spar will take place from the surrounding communities. There is evidence to suggest that many deep sea sediment species may be able to survive burial (under 5-6 cm of sediment) and full recovery of the benthic communities can occur within 12 - 24 months (Kukert and Smith, 1992). Thus, the long term effects on the sediment communities from the sea bed disturbance would be negligible. The buoy like other deep sea wrecks, will provide a new surface for fixed animals to grow on, and these in turn will provide a habitat and source of food for mobile animals such as worms, shrimps, crabs and deep sea fish (AURIS, 1994). .
6.3. LONG TERM FATE OF THE STRUCTURE ON THE SEA BED
6.3.1. The bulk of the Brent Spar is composed of structural steel (approximately 7,600t) and a solid concrete and haematite ballast (6,800 t). The long term fate of the main hull of the buoy will be largely independent of the state in which the buoy arrives at the sea bed (e.g. whether the tanks are ruptured), but will depend on the rate of corrosion of steel in the deep ocean. At the time of disposal, the steel structure is protected by a sacrificial anode system which is expected to be exhausted within 15 years. Studies have indicated that steel corrodes very slowly in deep water and anticipated corrosion rates at 4degC would be in the region of 5mm in 1000 years (Metocean, 1993). Assuming a typical hull thickness of 20mm, the structure is likely to remain largely intact for up to 4,000 years.
6.3.2. Other metallic components of the structure (e.g. copper wire) will also slowly corrode. The concrete/haematite ballast and any wood and plastic within the structure will remain largely intact. Some degradation, leading to mechanical breakdown of these materials would be expected with long term 'weathering' of concrete, biodegradation of wood and leaching of plasticisers from plastics.
6.3.3. Eventually, the hull and contents will collapse, and residual debris will be buried by sedimenting material. Sedimentation rates vary from 0.1-0.2 cm/1000 years in areas of low surface productivity to 20-30 cm/1000 years on the continental slope (Gage and Tyler, 1991). Assuming a sedimentation rate of around 20 cm/1000 years, once the structure has collapsed onto the sea bed, it will take several thousands of years to bury and become part of the deep sea sediment system.
6.4. ENVIRONMENTAL IMPACTS FROM THE PHYSICAL PRESENCE OF THE STRUCTURE
6.4.1. It is apparent that the structure will remain on the sea bed for several thousands of years before it eventually breaks up into debris and becomes incorporated into the sediment system. Apart from the colonisation of the structure by organisms and the formation of an 'artificial reef', the other potential effects from the physical presence of the structure will be possible impacts on other resource users. This possibility was considered during the site selection process, and the sites eventually chosen satisfied this criteria. The water depth is sufficient to preclude commercial fishing activities, and the proposed locations avoid main shipping routes, amenity areas, military activity, sub-sea cable routes, other disposal sites and prospective areas for oil, gas or minerals.
7.1. BIOLOGICAL EFFECTS OF HYDROCARBONS AND HEAVY METALS 7.1.1. Hydrocarbons and metals are widespread in the oceans, arising from geological, biogenic and anthropogenic sources. The concentration ranges found in nature in sea water and marine sediments are illustrated. Harmful effects on marine organisms may occur if they are exposed to concentrations of metals or hydrocarbons higher than those found in their normal environment. .
Note: 1. Neff, Rabalais and Boesch 1987;2. Metocean 1993; 3. Neff and Anderson 1981; 4. Clark 1993.
7.1.2. The intake of hydrocarbons is not essential to any organisms (except certain species of fungi and bacteria) to sustain their metabolism, although they may be synthesised by some species for various functions (e.g. protective wax on leaves of higher plants, buoyancy in marine invertebrates). Exposure to petrogenic hydrocarbons can result in a direct physical effect and/or acute or chronic toxic effects, depending upon the quantities encountered. Physical effects are typically seen after the bulk release of oil (cf. slicks occur at concentrations of > 15ppm in still water) when it coats the organism or its environment. The presence of this layer of oil may cause the loss of buoyancy and heat insulation of birds or marine mammals and asphyxiation in invertebrates, fish, birds and marine mammals. Specific components of crude oil can also produce toxic effects. The most potent in this respect are the aromatic and polycyclic aromatic compounds which can represent some 15-20% (by weight) of the oil (Neff and Anderson, 1981). The effects are species dependent; lethal concentrations to fish and marine invertebrates range from 1-500 ppm and sub-lethal concentrations occur at > 1ppb (Capuzzo and Kester, 1987).
7.1.3. Certain metals are essential to the metabolic processes of living organisms. Metals, in particular the heavy metals (i.e. transitional metals + metalloids) can also exert toxic effects. Transition metals (e.g. iron, copper and: manganese) are essential in low concentrations but can be toxic at high concentrations. Metalloids (e.g. mercury, cadmium, chromium and lead) are generally not required for metabolic activity and can be toxic at low concentrations (Clark, 1993). Concentrations of heavy metals that are considered acceptable to protect the marine environment are defined in Environmental Quality Standards (EQS) introduced under the EEC Directive 76/464/EEC. EQS values for key heavy metals (dissolved) in marine waters are Cadmium - 2.5 microg/l, Mercury - 0.3 microg/l, Lead - 15 microg/l. Chromium - 15 microg/l, Copper - 5 microg/l and Zinc - 40 microg/l (Metocean, 1993; AURIS 1994).
7.1.4. Marine organisms exhibit a range of strategies for dealing with elevated concentrations of hydrocarbons and heavy metals. These vary inter- and intra-specifically depending on the specific metal or hydrocarbon encountered. Some species are capable of partial or complete metal or hydrocarbon regulation, while others resort to accumulation strategies or a combination of both (Boehm, 1987; Clark, 1993; Capuzzo and Kester, 1987; Waldichuk,1985; Neff, 1979). Some species are able to change the rate of excretion to match the increase in uptake, for example zinc regulation in Palaemon elegans (Rainbow et al., 1990) and regulation of polynuclear aromatic hydrocarbons in fish, polychaetes and shrimps (Neff and Anderson, 1981; Capuzzo and Kester, 1987). Some marine organisms are able to concentrate contaminants in particular parts of their body. For example, high concentrations of zinc in the jaws of nereids and the jaws of Glycera, and copper accumulation in the gills of Melinna palmara (Clark, 1993). Bivalve molluscs tend to be poor regulators of metals and hydrocarbons and tend to accumulate them when exposed to elevated concentrations. However, when removed from the source of contamination these animals are often able to release the contaminants back to the environment (Neff and Anderson, 1981; Capuzzo and Kester, 1987; Clark, 1993).
7.1.5. Fundamental to producing an adverse effect, the heavy metals and hydrocarbons must be in a form that is biologically available. In the case of metals, this is usually the dissolved ionic form. With hydrocarbons the dissolved or very finely dispersed fractions are generally considered to be 'available'. Metals and hydrocarbons in particulate form in the water column or sediments are seldom directly available to organisms (Boehm, 1987; Waldichuk, 1985). Small amounts may be released into the gut if the particles are ingested, although evidence from deposit feeders suggests that availability through this route is limited (Neff and Anderson, 1981; Waldichuk, 198S). Similarly, metals in their pure form are unavailable to organisms unless they corrode and their corrosion products are soluble in sea water.
7.1.6. The association of heavy metals and hydrocarbons with particulate matter and sediments is well established. Apart from the lower molecular weight aromatic fraction, petrogenic hydrocarbons are generally hydrophobic and tend to associate with particulate material (Boehm, 1987; Mackay and McAuliffe 1988). The association of heavy metals is more complex and depends on factors such as pH, ionic strength and redox state (Salomons, 1992). Of particular importance are the effects of pH; at a low pH metals tend to go into solution and at a higher pH they tend to become associated with particulates. Thus in the open sea, metals will tend to be scavenged from the water by particulate matter.
7.1.7. Metals and hydrocarbons associated with particles entering the sediments are generally considered to be fixed, or firmly bound, unless certain physical (e.g. sediment resuspension), biological (e.g. biodegradation, bioturbation) or anthropogenic (e.g. dumping, aggregate extraction) processes result in their mobilisation. Generally, the upper sediment layers are well mixed to a depth of c. 8 - 10 cm by bioturbation (Smith, 1992). Within this layer, the organics, including hydrocarbons will be subject to aerobic biodegradation. In deeper sediments, where the oxygen demand is greater than supply, a reducing environment exists and anaerobic decomposition of organic material occurs.
7.1.8. The availability of heavy metals and hydrocarbons in sediments will be affected by reduction-oxidation (redox) processes. As oxygen levels decrease, biodegradation of hydrocarbons will effectively cease and unless the sediments are disturbed, they will remain associated with buried sediments. Similarly, most metals will be changed from their oxidised to reduced state (generally insoluble sulphides) and unless exposed to oxygen through some form of disturbance will remain fixed. Iron and manganese are the exception; the sulphides of these metals are more soluble than the oxides/hydroxides. Consequently, iron and manganese will be mobilised into the pore water and diffuse to the upper oxygenated layer where they will be re-deposited as oxides. If the whole sediment is anoxic, then these metals can diffuse and be re-mobilised into the water column. Although manganese and iron are not particularly toxic, their oxyhydroxides can scavenge metals such as copper, zinc and lead from the water. This, together with anaerobic formation of organo-metallic compounds such as tin, mercury and lead, provides a mechanism for making them biologically available (Waldichuk, 1985; Boehm, 1987; Salomons, 1992).
7.2. SLOW RELEASE OF CONTAMINANTS INTO BOTTOM WATERS
7.2.1. In this scenario, it is assumed that the buoy reaches the sea bed intact and sustains minor damage to one or more of the storage tanks upon impact. It is assumed that the release of contaminated water and sludge starts immediately from small breaches in the tanks and continues at a uniform rate over a period of c. 1000 years, as the remainder of the structure breaks-up. These assumptions are believed to represent a worst case for this scenario, as the predicted corrosion rates indicate that structure may take longer to break-up. It is also quite possible that the sludge may never be released and simply become incorporated into the mass of the disintegrating structure.
7.2.2. Under these conditions, release rates of residual contaminants in the storage tanks would be approximately 0.298 mg/sec - hydrocarbon, 0.00136 mg/sec - copper and 0.00277 mg/sec zinc in the sludge and approximately 0.06 mg/sec - hydrocarbon, 0.028 mg/sec - aluminium and 0.018 mg/sec - zinc in the contaminated water. The other main source of residual contaminants on the structure are heavy metals contained as pure metals or alloys within the structure and equipment. These would be expected to slowly leach as the metals corroded. Of all the metals on the structure, the sacrificial anodes will corrode the fastest, with estimated release rates of 60 mg/sec - aluminium and 19 mg/sec - zinc (Metocean, 1993). Thus, the buoy would act as a very low level chronic point source of hydrocarbons and heavy metals.
7.2.3. Once released to the environment, dissolved materials will be dispersed by the prevailing currents. A proportion of the particulate-bound materials will probably be picked up by the prevailing currents and dispersed; the remainder will settle to the sea bed sediments in the immediate vicinity of the point of release, although some could then be relocated as sea bed load.
7.2.4. The dissolved metals will either associate with particulate material in the water column and enter the sediment system, or disperse to background concentrations in the water column. Estimates of near field dispersion of zinc and aluminium from the anodes indicated that background levels would occur within 100m of the structure (Metocean, 1993). Leakage rates of aluminium and zinc from the storage tanks and corrosion rates of other metallic components would be several orders of magnitude less than the anodes and a similar level of dilution would be expected much closer to the structure. Simple modelling of dispersion of dissolved of finely dispersed hydrocarbons, indicate that background concentrations will be reached within 100m of the structure. Thus, any effects would be limited to organisms on or immediately adjacent to the structure. Evidence from fouling studies on oil and gas installations in the North Sea indicate no adverse effects to organisms living in the immediate vicinity and on sacrificial anodes (AURIS, 1994). This would indicate that any effects within the mixing zone (i.e. between the release point and reaching background concentrations) will be insignificant.
7.2.5. Particulate material will settle to the sea bed at a rate proportional to its particle size and density and become incorporated into the sea bed sediments. The metals and hydrocarbons associated with the sludge would be expected to be tightly bound to the particles and not be immediately available to marine organisms. It is believed that the majority of the sludge is composed of silt and sand which would be expected to quickly settle to the sea bed, immediately adjacent to the point of release. With settling velocities between 0.01-0.07 m/s, current velocities of 10 cm/s and a release height of 15 m, the particulate material would initially be expected to impact the sea bed within 150 m of the release point. Assuming the particles settle with an even distribution within 150 m radius of the point source (i.e. an area of 70,695 m2), a sedimentation rate of 0.2 mm/y and no sediment mixing, re-suspension or degradation processes, the level of hydrocarbons in surface sediments within this localised area would increase by up to two orders of magnitude. With the exception of cadmium which may increase marginally, elevations of the heavy metals would be negligible above sediment background levels.
7.2.6. Sediments contaminated with particulate bound material will be available for re-suspension, redistribution as sea bed load, biodegradation and burial as a result of bioturbation or sedimentation (Gage and Tyler, 1991, Gardener and Richardson.-1992; Smith, 1992; Boehm, 1987). The net effect of these processes would be to redistribute material into the upper layer of sediment and possibly move some of the contaminated material outside the primary deposition area. To provide an indication on the possible accumulation rates a simple box model approach has been used (e.g. Walsh, 1992). Assuming all the material remains in the primary deposition zone (i.e. is uniformly distributed within a 150 m radius of the structure), and that there are no losses, only dilution through bioturbation in the upper 10 cm, this would result in a gradual accumulation of sediment hydrocarbons at a rate of some 20-30% of the background per year, until the source of contamination was exhausted. In reality, most of the particulate-bound materials will be concentrated in the immediate vicinity of the point source and decrease in concentration away from the structure and it is unlikely that elevated sediment hydrocarbons would be detectable much beyond 500m of the structure. Detectable effects on benthic communities have been observed at increases in sediment concentrations of 10 - 60 mg/kg, (Davies and Kingston, 1992). Thus, in this scenario, biological effects would be expected within the immediate vicinity of the structure, probably within 150 - 250 m. These would probably be manifested by a gradual reduction in species diversity and appearance of more opportunist species.
7.2.7. It should be noted that these are conservative assumptions as factors such as re-suspension, biodegradation and burial as a result of sedimentation have been excluded. Biodegradation of the hydrocarbon fraction will occur under aerobic conditions. Data suggests that biodegradation of petrogenic hydrocarbons in sediments is a long term process, occurring on a time scale of years (Rartha and Atlas, 1987). For example, data from oil based cuttings piles in the North Sea, suggest degradation rates in surface sediments of some 25% in 18 months (Mair et al., 1987). Re-suspension may also be an important mechanism for redistributing surface sediments, particularly during periods of 'benthic storms'. It has been suggested that this process can re-mobilise some 10 - 60 % of the biogenic component of the sediment (Walsh, 1992). Eventually, materials entering the sediments will become buried. Typically it has been estimated that some 5% of the biogenic material entering sediments is buried (Walsh, 1992). Each of these processes would serve to further disperse or dilute contaminants, and in the case of hydrocarbons, degrade the small quantities of contaminants present. Taking full account of the likely mode of release of materials contained within the structure, and the factors referred to above, it is unlikely that particulate-bound materials will have any significant impact on the marine sediments outwith the immediate vicinity of the abandoned structure.
7.2.8. The residual hydrocarbon waxes lining the storage tanks would be expected to be fairly inert and resilient to any form of degradation in the anaerobic environment of the storage tanks. Once the structure started to disintegrate through corrosion (several thousand years), then aerobic biodegradation of any residual wax may occur. More likely, the wax will disintegrate with the structure and eventually become incorporated into the sediment system.
7.3. RAPID RELEASE OF CONTAMINANTS INTO BOTTOM WATERS
7.3.1. In this scenario, it assumes that the storage tanks rupture as the buoy hits the sea bed, releasing the water and sludge over a period of c. 24 hours. Apart from the effects of the rapid release of contaminated water and sludge, the environmental impacts will be the same as those described in the slow release scenario (see Section 7.3). Again, this assumption is believed to represent a worst case for this scenario, as it is highly unlikely that such a rapid release would occur.
7.3.2. The contaminated water will have equilibrated at the temperature of the North Sea bottom waters, which would be around c. 7degC during the summer. Thus, the release of this water at the sea bed, with temperatures of c. 4degC, will result in a plume that moves with the ambient current and is also slightly buoyant. The plume will pass through the sediment cloud raised by the impact of the structure with the sea bed and mix with the ambient sea water. Mixing with higher pH bottom water will encourage heavy metals in solution to adsorb onto particulate material in the water column. Any remaining soluble hydrocarbons and heavy metals will disperse as the plume slowly moves laterally and vertically. Given the small density differential between the plume and ambient water, the plume would not be expected to rise more than a few tens of metres from the sea bed, but simple dispersion modelling indicates that it could move several hundreds of metres laterally before completely dispersing.
7.3.3. As the plume moves and disperses, it will inevitably encounter, or be encountered, by bathypelagic fish inhabiting the area. Given the low densities of species living at these depths (c. 1 per 2,000 m2 ) and the observation that fish appear to be able to avoid areas of contamination, encounter rates with the plume will be low and of a short duration. Thus, toxicity thresholds are unlikely to be exceeded and the consequent effects will be negligible.
7 3.4. The sludge would drift with the current and quickly settle to the sea bed. Over the relatively short period of release, it is conservatively assumed that the current will flow from a single quadrant at c. 10 cm/sec. As before, the sludge would initially be expected to settle within some 150 m of the structure, some of it mixing with the sinking sediment cloud displaced as the structure hit the sea bed. Thus, the sludge would initially cover an area of approximately 17,500 m2, equivalent to loads of some 532 g/m2 - hydrocarbon, 2.43 g/m2 of copper, 0.33 g/m2 of cadmium, 0.51 g/m2 - of lead and 4.95 g/m2 - of zinc being deposited on the sediments, within the 24 hour release period. Because of the rapid rate of release, it would be expected that a significant proportion of the contaminated particulate material would be re-suspended and moved further afield with the sediment load.
7.3.5. In many ways, this scenario would be similar to a small one-off discharge of oil based mud and cuttings at an exploration well site and the effects on the benthic communities would be expected to be similar (e.g. Davies et al., 1984; Boehm, 1987; Yunker et al., 1990). High levels of hydrocarbons (up-to three orders of magnitude above background) and to a much lesser extent some of the heavy metals would be found within the primary deposition area, with concentrations decreasing away from the structure. Measurable levels of hydrocarbons may be detectable in the sediment up to 1000 m or so from the structure. Biological effects on sediment organisms would be expected to be more localised, with few observable effects beyond c.500 m or so. Immediately adjacent to the structure, the benthic fauna would be highly disrupted by both smothering (from the sludge and sediment displaced during impact of the structure on the sea bed) and any residual toxicity of the contaminants. Beyond the contaminated zone, which may extend c.150 - 250 m beyond the structure, the benthic communities may show reduced diversity, due both to residual contamination and to physical disruption of the sediments during impact of the structure. Beyond c. 500m, effects are likely to be undetectable and benthic community diversity similar to background.
7.3.6. Immediately after complete release of the contents of the tanks and deposition of the particulates on the sediments, the recovery process will start. Field studies indicate that affected areas in shallow coastal waters recovered within 3 months, whereas little recovery was apparent at a deeper water site within the same time frame (Yunker er al., 1990). At deeper water sites in the North Sea, there is evidence of the start of recovery in the surface sediments within a period of 18 months of ceasing drilling activities (Mair et al., 1987). It has been suggested that metabolic processes and recovery in the deep sea are slower than rates encountered on the continental shelf (e.g. Grassle and Morse-Porteus, 1987). However, more recent studies suggest that this might be an artefact of the methodology and the rates significantly higher than originally thought (Kukert and Smith, 1992). In comparison with continental shelf regions, the deep sea environment is stable, and apart from the occasional 'benthic storms', has few physical mechanisms to enhance dispersion or degradation of sediment contaminants.
7.4. RELEASE TO NEAR SURFACE WATERS
7.4.1. In this scenario, it assumes that one of the cracks in the damaged storage tanks opens shortly after the buoy leaves the sea surface (i.e. within 100-200m), releasing some 10% of the total water and sludge. Apart from the partial release of contaminated water and sludge near the surface, the environmental impacts once the structure arrives at the sea bed will be similar to one of the two scenarios outlined previously, and subject to the same probabilities of occurrence as specified in Section 6.
7.4.2.In this scenario, the buoy will be sinking vertically through the surface waters, creating an area, of highly turbulent flow around and behind the structure. Thus, any material ejected during descent will be vigorously dispersed vertically over a distance of a few hundred metres. Once in the water column, the relatively high current velocities found in the surface waters of the region (cf. in surface waters currents can be in excess of 75 cm/sec at the surface - Metocean, 1993) would be expected to rapidly disperse and dilute the soluble contaminants to background levels. The sludge would move with the currents and sink to the sea bed. Given the depth of water and typical sinking velocities of silt and sand, the particulates would be dispersed over a distance of several kilometres and effectively diluted to background levels.
7.4.3. Given the relatively small quantities of dissolved material that are considered likely to be released into the water column, and the horizontal and vertical dispersion of released materials, it is unlikely that elevated concentrations of contaminants would persist in the water column for more than a few minutes in the case of heavy metals, and a few hours in the case of the hydrocarbons. The vigorous mixing caused by the movement of the buoy would be expected to disperse any free oil into the water column in a similar way to dispersion and weathering of an oil slick (e.g. Mackay and McAuliffe, 1988). Thus, a surface sheen would not be expected. Potential exposures of pelagic organisms to elevated levels of heavy metals and hydrocarbons will therefore be low and the predicted impacts to such organisms would be negligible (AURIS, 1994).
7.5. ENVIRONMENTAL IMPACTS OF HYDROCARBON AND HEAVY METAL RELEASES
7.5.1. The deep ocean has long been considered for disposal of various wastes on the basis that it represents a vast area of unused resource (Angel, 1992; Pequegnat, 1983). Over 50% of the world's surface is located below the 3000 m depth contour. Apart from sub-sea cables, military activities, research, 'experimental mining' and limited dumping, the deep ocean is not utilised by man; currently commercial fishing operations do not extend to depths beyond about 1000 m, although experimental system are being developed to work to 1500 m depths.
7.5.2, In the first two scenarios (i.e. slow release and rapid release) highly localised effects on the environment are predicted. In the slow release scenario, elevated levels of hydrocarbon and heavy metals would be concentrated in the sediment close to the structure, although hydrocarbon may be detectable up to c.500 m and biological effects up to 150 - 250 m from the structure. The effects of the rapid release scenario, would follow a similar pattern although the area affected would be expected to be larger (hydrocarbon may be detectable up to c.l000 m and biological effects up to 500 m from the structure) but the effects would be over a much shorter duration. Given the small numbers of organisms present in the deep ocean and low encounter rates with the affected area, effects will be insignificant. Evidence from sites contaminated with oil based mud cuttings on the continental shelf suggests that there are indications of fish tainting (i.e. food chain effects) from fish caught close to sites where extensive drilling operations had occurred but not at sites with low levels of contamination (Davies and Kingston, 1992). Extrapolating these findings to the deep sea environment, this would suggest that food chain effects at the disposal site will not occur.
7.5.3. Below about 1500 m, the ocean is relatively isolated from the productive surface waters. The turnover time of water in the deep Atlantic is around 250 years. Primary production occurs in the surface waters and its downward flux provides the only source of food for the deep ocean. Apart from occasional movements of some species (e.g. deep sea squid) across 'the 1500 m contour', the deep ocean remains relatively isolated from surface waters. Thus, transfer of contamination from the deep sea system to the shallow water systems is not considered viable.
7.5.4. In the third scenario, involving partial release of contaminants close to the surface, additional effects would be observed in the water column. The relatively small quantities of material released, the vigorous initial mixing and good dispersion characteristics of the surface waters of the NE Atlantic would quickly reduce contaminant levels to background. No significant effects would be predicted for pelagic organisms within the 'sinking zone'.
8.1. APPROACH TO ASSESSMENT OF RADIOLOGICAL IMPACT
8.1.1. When released into the ocean, radio nuclides are subject to the same physical, chemical and biological dispersion, re-concentration and removal processes as stable isotopes of their corresponding elements; the only difference is that the concentrations of radio nuclides will in general decrease with time as a result of their radioactivity. Physical dispersion in the ocean waters occurs through current movements and mixing processes. In the ocean interior radio nuclides interact with organic and inorganic particulates which are moving down through the water column to the ocean floor. Some of these particulates wholly or partially dissolve during their descent, especially at depths below the carbonate saturation depth. releasing radio nuclides adsorbed on them back into the water column. Other particulates dissolve mainly at the ocean floor. At the floor, processes of bioturbation, diffusion in bed sediments, and burial in bed sediments, both return radio nuclides to the water column and remove them to the deeper sediments, which act as a permanent "sink". The degree to which particulate scavenging in the water column and interactions with bed sediments affect radio nuclide concentrations in ocean waters depends on the chemical properties of the relevant element. For example, thorium interacts strongly with particulates and bed sediments, radium much less strongly. In general, scavenging is more important at the ocean edges (i.e. in shelf seas and coastal waters) than in the open ocean, because particulate concentrations are higher closer to land.
8.1.2. Naturally occurring radio nuclides such as those in the Brent Spar sludge and scale are present in the oceans as a result of input from rivers, the atmosphere, and radioactive decay of their parent radio nuclides in the water column and in bed sediments. Cycling of natural radio nuclides in the ocean is used to study oceanographic processes and hence much is known about their behaviour (see for example, Guary et al., (eds) 1988; GESAMP, 1983; IAEA, 1988a; NEA, 1988; NEA, 1989 and references cited therein).
8.1.3. The most obvious way in which radio nuclides released into the ocean can cause radiation doses to humans is through consumption of seafood, particularly fish, molluscs, crustacea, and to a lesser extent seaweed. These organisms re-concentrate most radio nuclides: for example, the concentration of radium in fish (Bq per tonne) will be about 500 times the concentration in sea water (Bq per cubic metre). Other possible human exposure pathways include consumption of sea salt, consumption of desalinated water, inhalation of sea spray and suspended beach sediments, and external irradiation through boating, swimming and spending time on beaches.
8.1.4. It is assumed for radiological protection purposes that any radiation dose, however small, carries risks to human health, the risks of most concern being the induction of fatal cancer in the individual who receives the dose, and the induction of serious genetic defects in the individual's children. To assess radiological impact on humans it is therefore necessary to estimate potential doses, however small, rather than simply to show that doses will be below some "safe" level. Because radio nuclides released into the ocean disperse over long time scales and large volumes before reaching humans, dose estimates are made using mathematical models.
8.1.5. The disposal site for the Brent Spar has not yet been chosen, although investigations of potential sites are in progress. Furthermore, the quantities of radio nuclides in sludge and scale in the Brent Spar are relatively small (see Section 3.2). Thus, rather than carrying out detailed mathematical modelling, the approach used in this impact hypothesis is to use the results of internationally established models to scope potential doses. Details of the models and the ways in which their results have been used are given in the Annex. The estimated doses to human individuals for each scenario are given in Sections 8.2, 8.3, 8.4 and 8.5 and for populations in Section 8.6 below. Potential radiation effects on deep sea organisms are discussed in Section 8.7.
8.2. SLOW RELEASES OF RADIO NUCLIDES INTO BOTTOM WATERS
8.2.1. Two estimates have been made of the radiation doses which might be received by individuals in the scenario where the Brent Spar reaches the ocean floor intact and radio nuclides are released slowly into the ocean as the structure corrodes. One estimate is based on modelling for the 1985 review by the Nuclear Energy Agency of OECD of the continued suitability of the dumping site for radioactive wastes in the North East Atlantic (NEA, 1985). The other estimate is based on IAEA calculations for a generic ocean basin (IAEA, 1986). Details of the assumptions used in the two instances are given in the Annex.
8.2.2. Use of the NEA modelling results leads to an estimated maximum annual dose of per year to an individual who consumes above-average amounts of seafood. This dose is predicted to arise 500 years after disposal of the Brent Spar, and to be due to the consumption of fish and crustacea caught in the Antarctic Ocean. The radio nuclides contributing to the dose are radium-226 and its daughter products lead-210 and polonium 210. Doses from radium-228 and thorium-228 are predicted to be zero because they will decay before reaching waters where seafoods are harvested. The time and place of occurrence of the maximum annual dose are determined by the major features of water movement in the Atlantic and its neighbouring oceans. The NEA modelling is based on the view that deep water moves along isopycnal surfaces which, in the case of the Atlantic, outcrop at the poles. In this model the highest concentrations of radio nuclides in surface waters as a result of releases into bottom waters will occur in the Antarctic and Arctic Oceans. Concentrations tend to be higher in the Antarctic because they are made up of contributions from radio nuclides moving north, then west and south, from locations in the deep waters of the North East Atlantic, as well as contributions from radio nuclides moving directly south.
8.2.3. Use of the IAEA modelling results leads to an estimated maximum individual dose of 3 x10 -13 Sv per year to a seafood consumer, again from radium-226 and its daughter products. Doses from radium-228 and thorium-228 are estimated to be zero unless it is assumed that fish living in very deep ocean waters are caught and consumed, in which case the dose is estimated to be 1x 10 -14 Sv per year.
8.3. RAPID RELEASES OF RADIO NUCLIDES INTO BOTTOM WATERS
8.3.1. No specific estimates have been made of the doses which might be received if a fraction of the total activity present in the Brent Spar was released rapidly into bottom waters, as a result of damage to the structure caused by its impact on the ocean bed. Examination of the features of the models used by NEA and IAEA. shows that they would predict doses of the same order as those given in Section 8.2 for a rapid release, because in both cases doses depend more on rates and patterns of dispersion in the ocean than on rates of release into the ocean, provided the latter are rapid.
8.4. RELEASE TO NEAR-SURFACE WATERS
8.4.1. Dose estimates for a scenario in which some implosion occurs during the descent of the Brent Spar to the ocean bed have been made using an approach based on a model developed for a European Community project to evaluate the radiological impact of radioactivity in northern European marine waters (the MARINA project-CEC, 1990). Details of the calculational approach are given in the Annex. Assuming that 10% of the activity in sludge and scale is released during descent, and in the extreme case, that all of this activity reaches coastal waters to the west of Scotland, the maximum dose to a seafood consumer is estimated so be 9x10-9 Sv per yR
8.5. COMPARISONS WITH DOSE LIMIT AND EXEMPTION LEVEL
8.5.1. The significance of the doses estimated above needs to be judged in two ways: first to determine whether they are below the level which is regarded as intolerable, and second to determine whether they are potentially acceptable. The dose limit recommended by ICRP for members of the public is 1 mSv per year (10 -3 Sv per year) (ICRP, 1991). This limit is for all practices which can lead to radiation exposure of the public, except for medical procedures, and represents the level above which doses would be regarded as intolerable. The doses estimated above for the release to bottom waters scenario (Section 8.2) are less than one billionth of the ICRP limit, and the dose estimated for the release to near-surface waters scenario (Section 8.4) is less than one hundred thousandth of the limit. It is therefore clear that any doses arising from the ocean disposal of the Brent Spar would not remotely approach the limits of tolerability recommended by ICRP.
8.5.2. To judge the potential acceptability of the doses, the most relevant level is that given in an IAEA document on principles for exemption of radiation sources and practices from regulatory control (IAEA, 1988b). This document states that a radiation dose to an individual is likely to be regarded as trivial if it is of the order of 10 microSv per year (10 -5 Sv per year), and it is understood that IAEA will be using such a dose level in their work to define wastes which can be regarded as non-radioactive for the purposes of the London Dumping Convention. The doses estimated above for the release to bottom waters scenario (Section 8.2) are less than one ten millionth of the IAEA exemption level, while the dose for the release to near-surface waters scenario is less than one thousandth of the exemption level. It can therefore be concluded that doses to individual members of the public from disposal of the Brent Spar are well within the potentially acceptable range, and are well below the levels which have been suggested as de minimus for the purposes of regulatory control
8.5.3. For further perspective, the estimated doses can be compared to those which could be received from radio nuclides naturally present in sea water. The annual dose from such radio nuclides to a person who consumes above average amounts of seafood could be 2 - 5 mSv per year (2-5 x 10-3 Sv per year)(Pentreath 1988 and Annex to this report), that is a factor of over 200,000 greater than the highest individual dose estimated from disposal of the Brent Spar.
8.6. DOSES TO HUMAN POPULATIONS
8.6.1. Within the framework of radiological protection, estimates of doses to human populations ("collective doses") are used in determining compliance with the requirements that practices causing radiation exposure should be justified, and that all exposures should be as low as reasonably achievable, economic and social factors being taken into account (the ALARA or optimisation requirement) (ICRP, 1991). There are no limits on collective doses as there are for individual doses. However, the IAEA has suggested a level of collective dose to be used in determining whether practices might be exempted from regulatory control. This level is in terms of the collective dose which would be received by the whole exposed population, integrated over all time, from one year of carrying out the practice, and is 1man Sv (IAEA, 1988b).
8.6.2. In the case of radio nuclide releases into the oceans the population to be considered in collective dose calculations is that of the world. Clearly it is not possible to calculate the dose to every person in the world, for every year after the release, and sum them to obtain the collective dose. The procedure is to calculate radio nuclide concentrations in seafoods in each part of the ocean, as a function of time, using mathematical models. Seafood catch data for each part of the ocean are then used to estimate the total amounts of radio nuclides in seafoods potentially available for human consumption, and it is assumed that all the edible parts of the seafoods are consumed. This procedure has been used in studies with the NEA model, and in the European Community MARINA project, and the results are employed here to make very rough, pessimistic estimates of the collective doses to the world population, integrated over all time, from disposal of the Brent Spar (see Annex for details)
8.6.3. The estimated collective doses are 0.09 man Sv for release of all the activity in the Brent Spar to bottom waters, and 0.02 man Sv for release of 10% of the activity to near-surface waters. These values, which have been derived by methods which err on the side-of caution, are less than one tenth of the IAEA suggested exemption level of 1man Sv. They may also be placed in perspective by comparing them to the annual collective dose to the world population from natural radio nuclides in sea water of 300,000 man Sv (NEA, 1985). As with individual doses, it can be concluded that population doses from disposal of the Brent Spar are likely to be acceptable.
8.7. RADIOLOGICAL IMPACT ON DEEP SEA ORGANISMS
8.7.1. Organisms in the deep ocean are exposed to radiation from radio nuclides naturally present in sea water and sediments. The most straightforward way to assess potential radiation effects on organisms from the disposal of the Brent Spar is to determine whether, immediately around the structure, radio nuclide releases from sludge and scale will increase background radiation levels significantly.
8.7.2. The IAEA has developed a methodology for the calculation of doses to deep sea organisms from radio nuclide releases into bottom waters (IAEA, 1988a). Using this methodology, it is estimated that the maximum dose rate to deep sea organisms in the case of a slow release of radio nuclides from the Brent Spar would be about 3 nSv per hour (3 x 10 - 9 Sv per hour - see Annex). Natural background dose rates to these organisms are in the range 10 nSv per hour (10 -8 Sv per hour) to 5,000 nSv per hour (5x 10-6 Sv per hour) (NEA, 1985; IAEA 1988a). It can therefore be-concluded that in this instance it is highly unlikely that there will be radiation effects on deep sea organisms at the individual or species population level. More rapid releases would lead to dose rates higher than 3 nSv per hour (3x10-9 Sv per hour), but these would persist for shorter times and would be within the background range. Thus in these cases too, it is very unlikely that there will be effects on natural fauna.
8.8. ENVIRONMENTAL IMPACTS OF LSA SCALE RELEASES
8.8.1. The results from studies using internationally established mathematical models have been used to estimate the radiation doses which could be received by human individuals and populations from radio nuclides released from the Brent Spar.
8.8.2. The estimated doses to individuals are several orders of magnitude less than a level of dose defined by the IAEA as a basis for exemption of sources and practices from regulatory control. The estimated collective (population) doses are also well within an IAEA exemption level. These dose estimates are approximate but have been derived by methods which err generally on the side of caution. They are based on calculated average activity levels in sludge and scale; use of calculated maximum activity levels would not lead to doses significantly above those estimated.
8.8.3. It has also been shown that disposal of the Brent Spar is very unlikely to lead to radiation effects on deep sea organisms at either the individual species or the population levEL
9.1. The impacts of three sinking scenarios have been considered for the Brent Spar. The most likely scenario is that the buoy will sink to the sea bed and remain largely intact after impact. This would be followed by a slow release of the residual contaminants from the storage tanks to the deep ocean waters and sediments. The alternative of major damage to the storage tanks (and fast release of contaminants) on impact is considered less likely to occur. The third alternative of tank damage (and release of contaminants in the near surface environment) is considered unlikely to occur. The potential environmental impacts from each of these scenarios, is an extremely localised effect which, when considered in the context of the size of the ocean, will be of negligible significance.
9.2. The Impact Hypothesis for the most likely scenario is summarised in the following paragraphs. The buoy will sink vertically, due to the weight of the ballast in its base and hit the sea bed base first. The buoy would be expected to penetrate the surface layers of sediment (the extent depending on the precise characteristics of sea bed at the disposal site) and topple onto its side. The surge preceding impact, and the impact itself will displace surface sediments from the impact zone into the water column. The sediment cloud generated would move with the bottom currents and be re-deposited on the sea bed within 500 - 1000 m of the impact zone. Burial of the sea bed communities will be inevitable in the immediate area of impact. Smothering from the sediment cloud will also occur within 100 - 200 m of the impact zone. Beyond this area effects would be small.
9.3. After impact and resettlement of the sediment, re-colonisation of the disturbed sediments and new hard surfaces of the buoy itself will start from the surrounding communities. Recent evidence on recovery of disturbed sediments in the deep sea suggest that this process will take place relatively quickly, over a 12 - 24 month period.
9.4. The steel structure of the hull and other metallic components will also start to corrode at an estimated rate of some 5 mm/1000 years. This means that the structure is likely to remain largely intact for some 4,000 years. The concrete ballast, and any wood or plastic would remain largely intact although some degradation and mechanical breakdown may occur within these long time scales. Eventually, the structure would collapse onto the sediment surface and any residual debris eventually become buried under the slowly depositing sediments in 5 - 10,000 years.
9.5. The relatively small quantities of hydrocarbons and heavy metals remaining on the structure will slowly leach into the water column from the storage tanks and the remainder of the structure as it corrodes. Dissolved metals or hydrocarbons will disperse to background levels within c. 100 m of the structure. Contaminated particulate material from the storage tanks will initially be deposited within c. 150 m of the structure and enter the sediment system. Disregarding any processes which would subsequently disperse or degrade sediment contaminants, conservative estimates indicate that hydrocarbon and to a much lesser extent the heavy metal concentrations would be expected to slowly increase in the sediments immediately adjacent to the structure. It is unlikely that elevated sediment hydrocarbons would be detectable much beyond 500m of the structure. Detectable effects on benthic communities would be expected within the immediate vicinity of the structure, probably within 150 - 250 m. These would probably be manifested by a gradual reduction in species diversity and appearance of more opportunist species.
9.6. LSA scale is composed predominantly of barium sulphate, with smaller quantities of calcium and barium carbonates. In the deep ocean, some dissolution of the calcium salts would be expected. Thus, some of the naturally occurring radioactive elements may also go into solution, whilst the others would be expected to remain in the particulate phase. Conservative modelling of the long-term fate of the radioactive materials in the scale, indicate that the estimated dose rates to individuals (human) would be several orders of magnitude less than those regarded by the IAEA as de minimus for the purposes of regulatory control. Similarly, dose rates to human populations would be well within those acceptable to IAEA. The dose rates to deep sea organisms would be within the ranges encountered in nature and effects on natural populations would not be expected.
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A.1. RADIO NUCLIDE CONTENT OF SLUDGE AND SCALE
A.l.l. Most measurements of the radioactive content of sludge and scale have been to determine concentrations of the two radio nuclides which are of most concern in an operational context, namely radium-226 and actinium-228. For sludge in the Brent Spar storage tanks the measurements give an average radium-226 concentration of 4.5 Bq g-1, with a range of 1.2 to 8.0 Bq g -1, and an average actinium-228 concentration of 3.0 Bq g-l, with a range of 0.8 to 5.2 Bq g -1. Estimates of activities in hard scale in the Brent Spar have been made from records of on-shore disposals of scale from Brent Alpha and Brent Bravo. These give an average radium-226 concentration of 17.6 Bq g -l, with a range of 1.7 to 46.6 Bq g-1, and an average actinium-228 concentration of 15.2 Bq g -1, with a range of 1.3 to 50.8 Bq g -1.
A.1.2. For the radiological impact estimates it is necessary to know the concentrations of the long-lived radio nuclides, namely radium-226 and its daughter products lead-210 and polonium-210, and radium-228 (parent of actinium-228) and thorium-228 (daughter of actinium-228). Activities of lead-210, polonium-210, radium-228 and thorium-228 in sludge and scale are not measured routinely. However, there are results of gamma counting measurements available for lead-210, polonium-210 and thorium-228 in Brent Alpha scale/sludge. These results indicate that an assumption that radio nuclides are in secular equilibrium (i.e. that the activities of parent and daughter radio nuclides are equal) would be conservative. Hence, for the impact assessment it is assumed that concentrations of lead-210 and polonium-210 are equal to those of radium-226, and that concentrations of radium-228 and thorium-228 are equal to those of actinium-228. This assumption, together with assumed quantities of 100 tonnes sludge and 30 tonnes scale, leads to the estimates of activities in Brent Spar materials given in Section 3.2 of the main text.
A.2. MODELLING TO ESTIMATE THE RADIOLOGICAL IMPACT ON HUMANS OF RADIO NUCLIDE RELEASES INTO BOTTOM WATERS
A.2.1 Throughout the calculations, the activities in sludge and scale are considered together and it is assumed that radio nuclides are released from the Brent Spar into the ocean in soluble form, rather than as particulates which settle to the ocean floor. This assumption is pessimistic, particularly for hard scale.
A.2.2 For estimation of individual and collective doses from disposal of the Brent Spar use is made of the results of modelling studies for the disposal site for radioactive waste in the North East Atlantic (NEA 1985, Smith et al 1987). These studies employed a model for radio nuclide movement in the ocean which was developed by the Ministry of Agriculture, Fisheries and Food (MAFF) and the National Radiological Protection Board (NRPB) in the UK. Since the model was developed for the OECD Nuclear Energy Agency (NEA) review of the continued suitability of the North East Atlantic site, it is referred to here as the NEA model.
A.2.3 The model is of the compartment type and is one of the sorts of model recommended for use in GESAMP Report 19 (GESAMP, 1983). It consists of a model to predict rates and patterns of dispersion of radio nuclides by physical processes (advection and diffusion) and, overlaid on this, a model to represent radio nuclide interactions with particulates in the water column and with bed sediments (NEA, 1985). Apart from those representing waters immediately above the disposal site, the compartments in the model are defined to correspond with isopycnal surfaces (surfaces of constant density) in the oceans. This approach is used because water movement in the ocean is thought to occur primarily along, rather than across, these surfaces. There are 92 compartments in the model, of which 68 represent the Atlantic and 24 represent the Pacific. Antarctic, Indian and Arctic Oceans and the Mediterranean Sea The compartment structure of the model is shown in Figure Al; Figures A2 and A3 show the flow fields in the Atlantic which are represented through exchanges between compartments in the model. These exchanges were calculated from a general understanding of the oceanographic features of the Atlantic and knowledge of the existing distributions of conservative tracers (salinity and temperature). The structure of the model compartments around the disposal site is shown in Figure A4.
A.2.4 The sediment interaction part of the model is based on the work described in GESAMP Report 19. The model includes scavenging of radio nuclides by particles in the ocean interior, and bioturbation, diffusion and burial in sediments at the ocean boundaries (see Figure A5). The basic assumptions made in the model and its main features are as follows:
A.2.6 Interactions with bottom sediments are modelled by adding compartments to each of the bottom compartments of the dispersion model. These represent the sediment-water interface, the bioturbated zone, the sediment where mixing is due to pore water diffusion, and the deeper sediment which acts as a permanent sink for radio nuclides (see Figure A5). Interactions with suspended sediments in the ocean interior are modelled by including extra transfer coefficients between compartments of the dispersion model.
A.2.7 The model for radio nuclide movement provides estimates of radio nuclide concentrations In sea water in each region of the world's oceans, as a function of time after the start of release of radio nuclides into the oceans. Radiation doses to individuals are then calculated by considering those whose habits are such that they could incur higher doses than members of the population as a whole, via potential exposure pathways arising from disposal. Such groups are known as 'critical groups', not because they are at any particular risk but in the sense that so long as they are properly protected, the population as a whole will be safeguarded automatically. The exposure pathways considered in the NEA studies (and in IAEA work, see Section A.3) cover internal irradiation via ingestion and inhalation of radio nuclides, and external irradiation, and include both "actual" pathways (i.e. ones which could exist now) and "hypothetical" pathways (i.e. ones which might exist in future). The exposure pathways considered are as follows:
Actual pathways
Hypothetical pathways
* Consumption of plankton, consumption of deep water fish.
* For collective (population} doses only "actual" seafood consumption pathways are considered. Details of the parameter values used in the dose calculations are given in NEA (1985).
* The doses which could be received by individuals following disposal of the Brent Spar (see Section 8.2 - 8.4 of the main text) are derived by straightforward scaling of results given in NEA 1985 for disposal of radium-226. These results indicate a peak dose from radium-226 and its daughter products, summed over actual seafood consumption pathways, of 8.3 x 10-11 Sv per year per TBq of radium-226 disposed. The peak dose is predicted to arise 500 years after disposal, from seafoods caught in the Antarctic Ocean. It should be noted that this method of estimating the doses from disposal of the Brent Spar implicitly assumes that radio nuclides associated with the Spar initially will be released slowly over periods of the order of 1000 years. It also assumes that peak doses are not dependent on the location of the disposal site within the North East Atlantic. The first of these assumptions is not unreasonable. The second could lead to underestimation of Brent Spar doses, but it should be borne in mind that other aspects of the IAEA model are conservative, particularly the implied assumption that individuals actually eat substantial quantities of seafood caught in the Antarctic Ocean. In the Brent Spar dose calculations, no allowance is made for reductions in dose per unit intake factors for radio nuclides as a result of the 1990 recommendations of ICRP (ICRP, 1991; Phipps et al., 1991)
* The collective dose which could be received following disposal of the Brent Spar is derived using results given in Smith et al 1987 for sea dumping of unit activity of uranium-238. Only radium-226 is considered for the Brent Spar and it is assumed that consumption of fish is the dominant population exposure pathway. The collective dose incurred up to 104 years after dumping of uranium-238 is used, to make allowance for the difference in half-lives of radium-226 (1600 years) and uranium-238 (4.5 109 years). The calculations give two estimates of the collective dose to the world population, integrated over all time, from disposal of the Brent Spar: 0.09 man Sv and 5 x 10 -5 man Sv. Only the higher value is used in the main text (Section 8.6).
USE OF IAEA MODELLING RESULTS
A second estimate of the individual doses which could be incurred from slow releases of radio nuclides from the Brent Spar into deep ocean waters is made using the results of IAEA modelling (IAEA, 1984; IAEA, 1986). These models were developed by the IAEA in support of the London Dumping Convention and are described in Appendices VI and VII of GESAMP Report 19 (GESAMP, 1983) and deal with radio nuclide movement in a notional ocean basin.
The Appendix VII model is three-dimensional and is based on the assumption that physical dispersion in the ocean occurs only by diffusion. The model was developed primarily to predict the extent of the region of elevated contaminant concentrations around a source and will underestimate the degree of dispersion by physical processes unless sufficiently large eddy diffusion coefficients are chosen. The model includes interior scavenging (i.e. removal of radio nuclides from the water column by adsorption on to falling particulates) and boundary scavenging (i.e. burial by accumulating sediments, diffusion in sediment pore water and mixing within the bioturbated layer). Analytical solutions to the equations are available for the steady state situation (GESAMP, 1983; IAEA, 1984) and it is these which were used in the IAEA calculations (IAEA, 1986).
A.3.3 The Appendix VI model is one dimensional and is an equilibrium version of the more complex model described in Appendix IX of GESAMP Report 19. As well as diffusion. boundary scavenging and interior scavenging, it includes up welling of water and a corresponding downward movement. This feature means that the model is able to reproduCE fluxes of substances which are recycled between the surface and the deep ocean, and hence that the model is particularly useful for estimating concentrations of long-lived radio nuclides which are not very strongly adsorbed on particulates or bottom sediments. Despite its one-dimensional nature, the model provides adequate estimates of laterally averaged concentrations. As with the Appendix VII model, analytical solutions are available for the steady state situation (GESAMP, 1983; IAEA, 1984).
A.3.4 The IAEA used the Appendix VI and Appendix VII models to calculate the rates of release of radio nuclides into the ocean which, if continued for 1000 years, would lead to peak doses to individuals equal to the ICRP limit of 1 mSv per year (10 -3 Sv per year)(lAEA, 1986). The exposure pathways considered in these "release rate limit" calculations are the same as those in the NEA modelling (see Section A.2 above), as are the values of the parameters for these pathways. The sediment/water distribution coefficients (Kd values) and concentration factors for biological materials used in the IAEA calculations (IAEA, 1985) are, for almost all radio nuclides, the same as those in the NEA modelling (NEA, 1985). The release rate limits for each radio nuclide given by the IAEA are the lower of the values calculated with the Appendix VI and Appendix VII models.
A.3.5 For the five radio nuclides of interest in the Brent Spar situation the IAEA release rate limits are:
A.3.6 Assuming that the radio nuclides in Brent Spar sludge and scale are released at a uniform rate over a period of 1000 years, and scaling the release rate limits by the quantities of each radio nuclide in the Brent Spar, produces the dose estimates given in Sections 8.2 - 8.4 and 8.6 of the main text. The radio nuclide which gives rise to the highest dose is radium-226, for which the release rate limit is based on the Appendix VI model and is largely determined by the dose incurred from consumption of fish caught at a depth of 1500m. Since the Appendix VI model is one-dimensional and does not include movement of water (and hence radio nuclides) from one ocean basin to another it is not surprising that the Brent Spar dose derived from the IAEA release rate limits (3 x 10 -13 Sv y-1) is higher than that derived from NEA modelling (8 x 10 -14 Sv y -1). The volume and depth of the North East Atlantic are somewhat smaller than those of the notional ocean basin considered in the IAEA calculations (the volume and depth of the North East Atlantic are 5 x 10 16 m3 and 3.5 km respectively, while those of the IAEA notional ocean basin are 1 x 10 l7 m3 and 4 km). No allowance is made for these differences in deriving the Brent Spar doses because the factor involved (less than 2) is outweighed by the pessimism of the implicit assumption that fish caught at a depth of 1500m in the Antarctic are consumed in substantial quantities.
A.4. MODELLING TO ESTIMATE THE RADIOLOGICAL IMPACT ON HUMANS OF RADIO NUCLIDE RELEASES INTO NEAR-SURFACE WATERS
A.4.1 For the scenario in which some implosion is assumed to occur during the descent of the Brent Spar to the ocean floor, the approach used to estimate doses is based on parameter values in, and results from, a model developed for the Commission of the European Communities MARINA project (CEC, 1990: Charles et al., 1989). This model, known as MARIN1, is designed to estimate doses from liquid effluents discharged into north European waters from nuclear installations on land. The model is of the compartment type and its structure is shown in Figures A6 and A7.
A.4.2 To estimate individual doses it is assumed that 10% of the activity in Brent Spar sludge and scale is instantaneously released into near-surface waters above its disposal site, and that this activity moves rapidly into the area of sea represented by the "Scottish Waters" compartment in the MARIN1 model (see Figure A7). Peak concentrations in sea water where seafoods are caught are calculated by assuming that radio nuclides arc dispersed throughout Scottish Waters, without any removal by particulate scavenging or radioactive decay. It is then assumed that these peak concentrations persist for at least one year, and doses to persons consuming seafoods caught in Scottish Waters are calculated using the same parameter values as in the NEA and IAEA modelling (Section A.2).The Scottish Waters compartment of the MARINl model is relatively small and shallow (volume 1.3 x 10 13 m3, depth 110m) and current flows through it are such that the assumption that peak radio nuclide concentrations persist for a year is pessimistic (see exchange rates given in Figure A7). The total calculated individual dose of 9 x 10 -19 Sv y -l (see Sections 8.2 - 8.4 and 8.6) arises mainly from polonium-210 (7.5 x 10 -9 Sv y- l ) and lead-210 (1.3 x 10 -9 Sv y-l ),
A.4.3 To estimate collective doses for the Brent Spar scenario, use is made of MARIN1 collective dose results for unit discharge of uranium-238 from a land based site on the west coast of the British mainland (Charles et al., 1989), ratios of the collective doses from unit discharge of radium-226 and uranium-238 derived from Klos et al., 1989, and ratios of collective doses to the European and world populations given in Smith et al., 1987. From Charles et al, the collective dose to the European population, integrated over all time, from discharge of lBq of uranium-238 from Springfields is 1.9 x10-13 man Sv. From Smith et al., for unit discharge of uranium-238 into UK coastal waters about 73% of the total collective dose will be incurred by the European population and about 27% by the rest of the world population. From Klos et al., collective doses per unit discharge of radium-226 and uranium-238 are in the ratio of 890:1. Simple multiplication gives a collective dose of 0.02 man Sv to the world population for the Brent Spar scenario in which 9.3 x 10 7 Bq radium-226 is released.
A.5. MODELLING TO ESTIMATE RADIATION DOSES TO DEEP SEA ORGANISMS
A.5.1 As part of their work for the London Dumping Convention, the IAEA developed a methodology for the calculation of doses to deep sea organisms from releases of radio nuclides into bottom waters (IAEA, 1988). In this methodology the GESAMP Appendix VII and Appendix VI models (see Section A.3) are used to calculate radio nuclide concentrations in bottom waters and sediments, and simple dosimetric models are then used to obtain the dose rates to various types of marine organisms.
A.5.2 From Table A-III in IAEA (1988), the maximum dose rates to deep sea organisms per unit concentration of the Brent Spar radio nuclides in sea water are as follows.
A.5.3 Using the GESAMP Appendix VII model with the assumption that radio nuclides are released at a uniform rate over a period of 1000 years, and taking the "radius of the source" to be 70m (about half the height of the Brent Spar), the concentrations of these radio nuclides in bottom waters close to the Brent Spar are 4.3 x 10 -4 Bq m for radium-226, lead-210 and polonium-210, and 3.5 x 10 -4 Bq m-3 for radium-228 and thorium-228. Hence the maximum dose rate to deep sea organisms from all five radio nuclides is 3 nSv h-1. (3 x 10 -9 Sv per hour)
A.6. NATURAL BACKGROUND DOSES FROM RADIO NUCLIDES IN THE OCEAN
A.6.1 From IAEA (1988) natural background concentrations in sea water of the five radio nuclides of concern in the Brent Spar calculations are as follows (units are Bq m - 3):
A.6.2 Based on these concentrations and using the same parameter values as in the NEA and IAEA modelling (see Sections A.2 and A.3), the dose to a person who consumes above average quantities of seafood is calculated to be 0.3 - 5 mSv y-1 (0.3 - 5.0 x 10 -3 Sv per year). Most of this dose arises from polonium-210. The estimated dose of 2 mSv y-1 (2 x 10 -3 Sv per year) given in Pentreath (1988) lies within this range.
A.6.3 The collective dose to the world population from natural radio nuclides in the ocean, based on average radio nuclide concentrations and seafood catch data for the early 1980s, is estimated to be 3 x 10 5 man Sv per year (NEA, 1985).
A.6.4 Table IX of IAEA 1988 gives the range of natural background dose rates to deep sea organisms as 10 nSv h-l to 5 microSv h-l (l0-8 to 5 x 10 -6 Sv per hour).
CEC, 1990. The Radiological Exposure of the Population of the European Community from Radioactivity in the North East Atlantic. Proc. CEC Seminar, Bruges, June 1989. CEC, Luxembourg.
Charles, D et al, 1989. The Radiological Impact on EC Member States of Routine Discharges to North European Waters, Report of Working Group IV of CEC Project MARINA. National Radiological Protection Board Memorandum NRPB-M172. NRPB. Chilton. (Also published by CEC.)
GESAMP, 1983. IMO/FAO/UNESCO/WMO/WHO/IAEA/UN/UNEP Joint Group of Experts on the Scientific Aspects of Marine Pollution, Reports and Studies No.l9, An Oceanographic Model for the Dispersion of Wastes Disposed of in the Deep Sea. IAEA, Vienna.
IAEA, 1984. The Oceanographic and Radiological Basis for the Definition of High-Level Wastes Unsuitable for Dumping at Sea, IAEA Safety Series No.66. IAEA, Vienna.
IAEA, 1985. Sediment Kds and Concentration Factors for Radio nuclides in the Marine Environment, IAEA Technical Reports Series No.247. IAEA, Vienna.
IAEA, 1986. Definition and Recommendations for the Convention on the Prevention of Marine Pollution by the Dumping of Wastes and other Matter, 1972, 1986 Edition, IAEA Safety Series No. 78 IAEA Vienna.
IAEA, 1988. Assessing the Impact of Deep Sea Disposal of Low Level Radioactive Waste on Living Marine Resources, IAEA Technical Report Series No.288. IAEA, Vienna.
ICRP, 1991. 1990 Recommendations of the International Commission on Radiological Protection, ICRP Publication 60. Annals ICRP Vol.21 Nos 1-3.
Klos, R A et al, 1989. Calculations of the Radiological Impact of Unit Releases of Radio nuclides to the Biosphere from Solid Waste Disposal Facilities. National Radiological Protection Board Memorandum NRPB-M150. NRPB, Chilton.
NEA, 1985. Review of the Continued Suitability of the Dumping Site for Radioactive Waste in the North East Atlantic. NEA/OECD, Paris.
Pentreath, R 1, 1988. Radio nuclides in the Aquatic Environment. IN Radio nuclides in the Food Chain, edited by M W Carter et al, ILSI Monograph. Springer-Verlag, New York.
Phipps, A W et al, 1991. Committed Equivalent Organ Doses and Committed Effective Doses from Intakes of Radio nuclides. National Radiological Protection Board Report NRPB-R245. HMSO, London.
Smith, G M et al, 1987. Calculations of the Radiological Impact of Disposal of Unit Activity of Selected Radio nuclides, National Radiological Protection Board Report NRPB-R205. HMSO, London
LIST OF FIGURES