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Table 7 Calculated consumption emissions of lead in Sweden, 1880-1980
Source: Swedish Trade Statistics, various years.
Fig. 1 Calculated emissions originating from production and consumption of lead in Sweden, 1880-1980
When comparing different sources of lead emissions over time, the significance of automobiles and trucks is particularly striking. Over a period of less than four decades, automotive emissions of lead far exceed total industrial emissions for the whole century.
In order to describe the immission landscape at different times, a simple flow model was used, where emissions to land and water, flows from soil to water, and, finally, accumulation in sediments and soils were considered. Detailed maps of chromium and lead loads in Sweden from 1880 to 1980 have been published elsewhere (Bergbäck et al., 1989, 1992). In figures 2 and 3 a comparison is made for accumulated amounts of chromium and lead in soils and sediments between 1950 and 1980.
Fig. 2 Calculated amounts of chromium in soil and sediment in Sweden, 1950 and 1980
Fig. 3 Calculated amounts of lead in soil and sediment in Sweden, 1950 and 1980
These calculated amounts of metals in soils in Sweden correspond to measurements taken during the 1980s, particularly in areas where high production emissions had taken place. Exact comparisons, however, are difficult to make, as our calculations represent mean loads in administratively defined areas. Comparisons with monitoring data reveal similarities but also discrepancies in the patterns, and the latter could in some cases be explained by the composition of the bedrock. (Uncertainties in the calculations are further discussed in Bergbäck et al., 1989, 1992.)
The present rate of lead consumption in Sweden is approximately 25,000 tonnes per year, excluding ammunition and gasoline . If this rate were to remain constant between 1980 and 2080, another 2.5 million tonnes would be added to the 2 million tonnes already accumulated in the last 100 years, giving a total of 4.5 million tonnes of lead in the anthroposphere.
It would also be relevant to compare the anthropogenic release of lead into the environment with the mobilization of lead from the bedrock during the weathering cycle. Weathering mobilization may be calculated by using average trace metal concentrations in soils and the suspended sediment flux in rivers. According to Nriagu, the global weathering rate for lead is approximately 180,000 tonnes per year (Nriagu, 1990). In Sweden's case, this would mean about 500 tonnes per year. Thus, 50,000 tonnes might have been released by natural weathering processes within the last 100 years. This amount should be compared with our calculations of total emissions, which are approximately four times higher.
The rate of emissions (see figure 4) and accumulation of chromium is higher than for lead. The use of chromium products in Sweden (e.g. stainless steel) has increased dramatically since the Second World War. As long as consumption emissions remain at a high level, chromium will have a strong impact on the environment in the future.
In figure 5 the emission rates of lead, chromium and cadmium are compared, with a roughly calculated weathering rate for these metals. Obviously, the anthropogenic contribution is significant, especially for lead.
Even though production emissions in Sweden have decreased during the last few decades, the accumulation of lead and chromium in soils and sediments will continue owing to the dissipative consumption losses of various products. To give an example: Suppose consumption emissions remain on the 19701980 level while production emissions are assumed to be low or even negligible; then the calculated amounts of chromium in the soils of some urban areas (e.g. Stockholm) will be as high as they are in the most polluted industrial regions today within only a few decades (Bergbäck et al., 1989). Thus, urban environments can be regarded as ecological "hot spots" for toxic metals. Also, in the future agricultural soils in suburbanized areas may be close to exceeding their carrying capacity for trace metal pollution.
Fig. 4 Chromium and lead emissions in Sweden compared with supply (imports-export + production) in thousands of tonnes per year
The changing spatial pattern of heavy metal loads in Sweden reflects the dynamics of industrialization. The first industrial revolution was based on local resources, such as raw materials and energy sources. Later, with greater affluence and mobility, an "urban world" developed. Consequently, the pollution load in soils and sediments has altered from being a "defined pollution problem" within certain industrial regions to a situation where the end-use of products, together with the mobility pattern of goods, define the pollution problem.
Fig. 5 The ratio between total emissions and weathering for lead, chromium, and cadmium in Sweden
In a general sense, our results illustrate a new dimension of the landscape. Industrial and urban areas often have soils and sediments with a higher recognized level of heavy metals. In these areas the "societal weathering rate" exceeds the natural one. In rural areas, with a more natural background dominated by the average bedrock composition, the pollution load of heavy metals is still less pronounced. The consequences of this development are difficult to predict, but it is obvious that a new dimension will be added to the conceptualization of the landscape, with particular implications for land-use planning.
Anderberg, S., B. Bergbäck, and U. Lohm. 1989. "Flow and Distribution of Chromium in the Swedish Environment: A New Approach to Studying Environmental Pollution." Ambio 18:216220.
- 1990 Pattern of Lead Emissions in Sweden 1880-1980. Report 13/90. Swedish National Chemicals Inspectorate.
Ayres, R. U. 1978. Resources, Environment and Economics. New York: John Wiley & Sons.
Ayres, R. U., and A. V. Kneese. 1969. "Production, Consumption and Externalities." American Economic Review 59, no. 3: 282-296.
Ayres, R. U., and S. Rod. 1986. "An Historical Reconstruction of Pollutant Levels in the Hudson-Raritan Basin." Environment 28:14-43.
Bergbäck, B. 1992. "Industrial Metabolism. The Emerging Immission Landscape of Heavy Metal Immission in Sweden." Dissertation. Linköping Studies in Arts and Science, no. 76.
Bergbäck, B., S. Anderberg, and U. Lohm. 1989. "A Reconstruction of Emissions, Flow and Accumulation of Chromium in Sweden 1920-1980." Water, Air, and Soil Pollution 48:391407.
- 1992. "Lead Load: The Historical Pattern of Lead Use in Sweden." Ambio 21.
Nriagu, J. O. 199(). "Global Metal Pollution. Poisoning the Biosphere." Environment 32:7-33.
Tarr, J. A., and R. U. Ayres, 1990. In: B. L. Turner et al., eds., The Earth as Transformed by Human Action. New York: Cambridge University Press, p. 623.
William M. Stigliani and Stefan Anderberg
In November 1986 an accident at a pharmaceutical company located on the River Rhine at Basel, Switzerland, resulted in the inadvertent release of 33 tons of toxic materials.' The effects were immediate and dramatic: half a million fish and eels died and local residents could not use the river as a source of drinking water for about a month. This accident, highly publicized in the world press, raised a major public outcry calling for an action plan for reducing the risks of such accidents in the future.
Historically, however, the impact of chemical accidents on the overall pollutant load to the river has been relatively minor. For example, in 1980 about 27 tons of toxic materials daily (10,000 tons per year) were transported by the Rhine into the Dutch Delta and the North Sea. This toxic load was the result not of accidents but, rather, of normal industrial, commercial, agricultural, and urban activities conducted within the Rhine Basin on a routine basis.
The effects of such chronic pollution are not as obvious or spectacular as those occurring after industrial accidents. Much of the daily input ends up in sediments of the Dutch Delta, and the rest is washed out to the North Sea. Even today the sediments in the delta are so polluted that the spoils, collected during dredging operations to keep navigation lanes open, are too toxic to be applied to polder lands in the Netherlands, as was the practice in the past. On the other hand, the River Rhine today transports far fewer pollutants to the Netherlands than it did in 1980, even though the level of economic activities in the basin has not changed very much since then.
Analysing the history of pollution in the Rhine Basin, including the recent clean-up, can provide valuable insights into the linkages between economic activities and chemical pollution, and the opportunities for decoupling economic growth from environmental degradation. The research described in this chapter, while not addressing all possible aspects of this history, will, we hope, provide a basis for improved policy-making.
The Rhine Basin extends over five European nations (fig. 1). Included are most of Switzerland, the north-east corner of France, Luxembourg, most of the south-western Lander (provinces) of Germany, and most of the Netherlands. The population of the basin is about 50 million and the area is about 220,000 km². About half of the land is used for agriculture, one third is forests, and the remainder is urban and suburban areas.
The basin is perhaps the most heavily industrialized in the world. Although the stream flow of the Rhine comprises only about 0.2 per cent of the flow of all rivers, about 10 to 20 per cent of the total Western chemical industry (OECD countries) is located in its basin. Industry is particularly concentrated in the catchment areas of the Ruhr, Neckar, Main, and Saar tributaries. Little net sedimentation of heavy metals occurs until the flow reaches the Dutch Delta, which extends from the German-Dutch border to the North Sea. About 75 per cent of the metals are deposited in the sediments of the delta, and the remainder disperses into the North Sea.
Our study analyses the entire system by which resource inputs to the industrial economy are converted into outputs that must be absorbed and processed by the environment. For analysing a given chemical, this systems approach can be divided into three steps:
1. Identification of materials in which the
chemical is embodied, and the pathways by which they flow through
the industrial economy.
2. Estimation of emissions and deposition to air, water, and soil for each material at each stage of its life cycle.
3. Construction of a basin-wide pollution model for assessment of proposed emission reduction policies, environmental impacts, or other relevant issues related to the chemical in question.
Fig. 1 The Rhine Basin. Place-names in boxes signify locations of monitoring stations of the International Commission for the Protection of the Rhine
In step 1, it is essential not to miss any important source of pollution. In this regard, it should be noted that many chemicals enter the industrial economy inadvertently as trace impurities of high-volume raw materials such as fossil fuels and iron and non-ferrous ores. Moreover, a full accounting should be made of all stages of the material's life cycle. These include not only the stage of production, but also the later stages of use and disposal.
Overlooking important sources of emissions can be costly. For example, Tschinkel (1989) has noted that billions of dollars have been spent in the US on the construction of secondary sewage treatment plants. Many of the benefits gained from this technology, however, have been nullified because discharges of untreated storm waters containing toxic urban street dust continue to flow into lakes, rivers, and estuaries. Such an omission may not have happened had planners been more aware of the significance of street dust as a major source of toxic materials.
In step 2, emissions and deposition are estimated quantitatively. Emissions may be classified broadly into two categories: point source and diffuse.. Point sources include electric power plants, industry, incinerators, sewage treatment plants, and others. Their emissions are typically highly concentrated and confined to a specific location, usually within an urban area. For each type of point source, emission factors, generally expressed as weight of pollutant per unit weight of material consumed or produced, are assigned for emissions to air, water, and land. Emission factors may change over time, decreasing as cleaner technologies are implemented. Total emissions are calculated as the product of the emission factor and the weight of material consumed or produced.
Particularly in the case of atmospheric pollution, it is important to make a distinction between emissions and deposition (or immissions, as it is called in other languages). Via the mechanism of long-range atmospheric transport, emissions may be deposited hundreds or even thousands of kilometres from their sources. Thus, some emissions generated in the basin are transported and deposited outside the basin, and some emissions from outside the basin are transported into it. A long-range atmospheric deposition model has been developed at IIASA (Alcamo et al., 1992) for estimating deposition in the basin.
In contrast to point-source emissions, diffuse emissions are generally less concentrated, more dispersed spatially, and dependent on land use, which can be broadly categorized as forests, agricultural lands (both arable and grassland), and urban areas. The only inputs to forested lands are assumed to be atmospheric deposition via long-range transport. Chemical inputs to agricultural soils include not only longrange atmospheric deposition, but also agrochemicals, manure, and sewage sludge. Diffuse emissions from these two land uses are determined using a runoff export model (Jolankai et al., 1991).
Transport of pollutants to surface and ground waters is much greater from agricultural lands than from forested areas. Transport occurs via storm runoff, erosion, and vertical seepage. The relevant parameters to be determined are the rates of applications of particular chemicals, expressed as weight per hectare, and the partition coefficient, which determines the fraction of chemical that is mobilized and transported and the fraction that remains bound in the soil. Even if only a small percentage of the chemicals is mobilized, total emissions can be significant because of the enormous chemical inputs and the large spatial coverage of agricultural lands.
Another important source of diffuse emissions is transport of pollutants from paved urban areas to surface waters. This occurs by the build-up of toxic materials in street dust during dry periods, and the washing out of the dust during storm events. The pathways by which the transport may occur are shown in figure 2. There are three main sources of toxic materials in urban dust: corrosion of building materials (particularly for heavy metals such as zinc, used in galvanizing and surface materials), exhausts and lyre wear from automobiles and other road vehicles (important for lead and zinc), and local and longrange atmospheric deposition (a dominant source of cadmium).
When storm sewers are separate from municipal sewers (path SSS in figure 2), the pollutants are transported directly to surface waters. Alternatively, storm sewers may be connected to municipal sewers that discharge to surface waters without treatment (path CSSW in figure 2), or they may be connected to municipal sewers in which the effluents are treated (path WWTP in figure 2). In sewage treatment plants with primary and secondary treatment, typically 50 per cent or more of input heavy metals are trapped in sewage sludges. Even when the storm sewers are connected to sewage treatment plants, however, some fraction of the polluted street dust may be transported to the river unabated if the volume of storm flow exceeds the flow capacity of the sewage treatment plant (path CSO in figure 2), which is often the case. Another important source of pollution, also indicated in the figure, is the atmospheric deposition on unpaved urban areas, with subsequent seepage to ground waters and transport to the river.
Fig. 2 Pathways by which pollutants in urban areas are transported to surface waters (Source: Behrendt and Boehme, 1992)
To calculate emissions from corrosion, it is necessary to determine rates of corrosion per unit surface of the corroded material and the total surface coverage of the material in question. For instance, rates of zinc corrosion are strongly linked to urban SO2 concentrations and will decrease over time as SO2 levels are lowered. The following equation (ECE, 1984) shows the empirical relationship between SO2 concentration and the rate of zinc corrosion from galvanized steel:
Y = 0.45* [SO2] + 0.7.
where Y = annual corrosion rate of galvanized steel (g/m²/yr), and [SO2] = concentration of SO2 in air (mg/m³)
Emissions from road traffic owing to tyre wear may be estimated by determining emission rates per vehicle km, and multiplying this rate by vehicle km per year and the number of vehicles per year.
Lead emissions from combustion of gasoline may be calculated by multiplying lead emitted per unit of gasoline burned and multiplying by annual gasoline consumption. The emissions are allocated spatially by apportioning them according to traffic density.
In urban areas, local atmospheric emissions are particularly important, since a significant fraction, typically around 10 per cent of the total emissions, are deposited within 10 to 20 kilometres of the source. Factors affecting the proportion of local to long-range emissions include smokestack height, velocity of the gases and particulate matter leaving the stack, meteorological conditions, and particle size of emitted pollutants. The IIASA study includes an analysis of trends in local emissions as affected by changes in the abovementioned factors (Hrehoruk et al., 1992).
Lastly, it is necessary to employ an urban hydrology model that estimates the fraction of street dust that flows to the river. Even though urban and suburban areas occupy only about 15 per cent of total surface area in the basin, their contribution to the total diffuse load of aqueous emissions is significantly higher. This is because of the prevalence of hard, impermeable surfaces in urban lands (typically around 33 per cent of total urban area), from which run-off and transport of pollutants can be as high as 90 per cent, compared to a maximum of about 25 per cent for agricultural lands (Ayres and Rod, 1986).
Completion of steps 1 and 2 for a given chemical provides a pollution model of the basin, including inputs and outputs for the chemical, its flows through the industrial economy over time, and its spatial allocation for each time period of interest. The model can be used for various purposes. For example, a historical analysis of pollution can provide information on changing trends in pollution sources. In the case of pollution in the River Rhine, the IIASA analysis indicates that since the mid-1970s diffuse sources of emissions of heavy metals have become increasingly important relative to point sources. Another useful application of the historical analysis is the possibility for estimating the cumulative build-up of toxic materials in soils and sediments. Currently, hardly any information exists on the rates of accumulation of toxic chemicals over wide spatial regions, or on the evaluation of resulting impacts to the environment and human health. (For a comprehensive discussion of cumulative chemical loading and potential environmental impacts see Stigliani, 1988, and Stigliani et al., 1991.)
The pollution model can also be used to test the effectiveness of proposed policy options for reducing emissions of toxic chemicals. Because the model is based on mass balance analysis, all material flows to air, water, and land within the basin must be taken into account. The model will thus expose options that would not reduce overall emissions in the basin, but, rather, would transfer them from one pollution pathway to another.
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