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A cost comparison of alternative policies for sulphur dioxide control
A cost comparison of alternative policies for sulphur dioxide control The case of the British power plant sector
Heinz Welsch
The paper is concerned with the costs of achieving overall reduction targets for sulphur dioxide (SO,) emissions of the British electricity sector under two alternative regulatory strategies: unifarm limits to SO, concentration in$ue gas and transferable emission permits. The implications of the SO, abatement technology for the allocation of abatement activities under the two types of regulation are analysed, and overall abatement costs are compared. An aggregate cost function is developed which may serve, together with transboundary SO2 transfer coeficients, as a basis for eficiently setting overall emission reduction targets for countries participating in a coordinated SO, control policy. Keywords: Airquality management; Emission abatement costs; cost-effectiveness analysis
In the past one and a half decades the problem of transboundary air pollution has received increasing attention. As a result of this, in 1977 the UN Economic Commission for Europe established the ‘Cooperative Programme for the Monitoring and Evaluation of Long-Range Transmission of Air Pollutants in Europe’ (EMEP). From the atmospheric model developed within EMEP (see United Nations Cl93 a matrix of transfer coefficients can be derived which relates depositions of sulphur dioxide (SO,) in 27 European countries to SO, emissions in any one of these countries. This matrix reveals that for most of the countries more than half of the SO* emissions are exported. Regarding depositions, in most cases more than half are of foreign origin. For some countries, in
The author is with the Institute of Energy Economics, University of Cologne, Albertus-Magnus-Platz, D-5000 K6ln 41, FR Germany. Numerical results presented in this paper are based on data partly compiled by I. Hoven, W. Schulz and P. H. Suding. Their contribution is gratefully acknowledged, while the sole responsibility for any errors remains with the author. Final manuscript
received 15 April 1988.
0140-9883/88/040287-11
$03.00 0
1988 Butterworth
particular the Scandinavian ones, the ‘import share’ is considerably higher. Due to these findings the view was widely accepted that SO, control strategies have to be internationally coordinated. In 1984 a number of European countries and Canada agreed on reducing their SO, emissions by a uniform rate of 30 % in 1993 based on 1980 figures. In 1983 the Commission of the European Communities submitted a proposal for a Council Directive (EC [S]) requiring that Member States reduce annual SO, emissions from large combustion plants by 60% in 1995 relative to their 1980 emissions. Although, from a political point of view, these agreements are an achievement, one may wonder whether such uniform rates of emission reduction are an efficient way of reducing the effects of sulphur deposition. Using depositions in certain target areas as an indicator of damage, Hordijk [9] shows that the same degree of damage reduction which results from flat rate emission abatement all over Europe can also be achieved by selectively lowering emissions of certain countries, requiring a smaller decrease in overall emissions. This provides evidence that neglecting differences in the marginal damage caused by emissions of different origins, as expressed by unit transfer coefficients, may give rise to inefficiencies. However, from an economic point of view, this is only one side of
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the coin, since for economic optimization differences in marginal damage have to be balanced against differences in the marginal costs of emission abatement. Hence, to set efficiently emission reduction rates for the individual countries involved in a coordinated deposition-oriented SO, control policy it is not only necessary to know the matrix of transfer coefficients but also the ‘macroeconomic’ SO, abatement cost functions of these countries. Moreover, to guarantee efficiency, these target rates must then be implemented in a cost-effective way. Regarding this issue, it is well-known that the traditional regulatory strategies, such as prescribing uniform emission limits for individual emitters, are not cost-effective. This paper relates to both the efficient setting and implementation of emission reduction targets; however, it focuses on the latter. Its main purpose is to provide a nationwide cost comparison of two regulatory strategies for SO, control: uniform limits to the concentration of SO, pollutants in flue gas and transferable emission permits. Such a comparison is important because, though the superiority of transferable permits is theoretically evident, it is not a priori clear whether the amount of potential cost savings would justify a move away from traditional regulatory practice. Moreover, since transferable permits lead to a least-cost allocation of abatement activities for any given overall abatement level, a nationwide cost function of SO, emission abatement is obtained as a by-product of analysing the costs of a transferable emission permit policy. As mentioned above, such a cost function, if available for all countries participating in a coordinated SO, control policy, could serve, together with the transfer matrix, as a basis for efficiently setting emission abatement rates for the countries, given prespecitied deposition reduction targets. In the present paper the analysis is only carried out for one country as a leading example, namely the UK. For a more comprehensive analysis which covers all countries of Western Europe and their atmospheric linkages see Welsch [22]. Moreover, the scope of this study is restricted to the power plant sector. The reasons for this choice are as follows. According to the results obtained within EMEP, the UK is the largest producer of SO, in western Europe, emitting a total of 4.25 x lo6 t in 1982. Due to the policy currently pursued in the UK of alleviating local effects of SO, emissions by means of tall stacksand the prevalence of south-westerly winds, 63.4% of the total emissions or 2.7 x 1Oht are deposited outside its boundaries, mostly in continental Europe. Thus, in order to reduce depositions on the continent, the UK is a prime canditate for substantial reduction of its SO, emis-
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sions, and an anaklysis of the costs of emission abatement in the UK is particularly called for, taking into consideration alternative regulatory strategies for implementing overall emission reduction targets. Since the cost comparison of such policies relies on modelling the behaviour of individual emitters, it is necessary to keep their number in a manageable range and to make sure that data on individual emitters are available. Both requirements are fulfilled by the power plant sector. Taking into account that in 1982 it accounted for 65.6% of SO, emissions in the UK, it seems justifiable to confine oneself to that industry. This also has the advantage that there exist several previous studies on SO, abatement options in the British electricity sector, which may serve as a basis to build upon and, partly, as a standard for checking results. In particular, the work of Highton and Webb [6,7,8] has to be mentioned. However, their research focuses on the cost-effective choice of abatement technologies rather than on the assessment of regulatory strategies, and the overall reduction rates they consider are only in the range of up to 40% by 1995. Nevertheless, their results are valuable for the present investigation because, a priori, one might think that the type of regulation influences the choice of technology and the latter has to be accounted for in a cost evaluation of the former. But, following Highton and Webb [7,8], the abatement potential of technologies such as coal or oil desulphurization is rather limited, and flue gas desulphurization (FGD) seems to be the only suitable method to achieve reduction rates higher than about 20%. Therefore, given the reduction rates currently envisaged, the following cost comparison of regulatory strategies will be based on one technology only, namely FGD. It may be noted that cost comparisons of the kind carried out in this paper are not a new type of undertaking. There have been numerous such studies relating to the US bubble policy. A classic example is Atkinson and Lewis [ 11, and a more recent one Seskin, Anderson and Reid [ 161. However, to the author’s knowledge, no such analysis has been performed on a nationwide scale, as is necessary for any attempt at an efficient, coordinated control of transboundary air pollution.
Sulphur dioxide emissions in the British electricity sector The projections of annual SO, emissions refer to the early 1990s. They are based on a power plant inventory comprising coal- and oil-fired plants with a capacity of 1OOMW or more, the majority being operated by the Central Electricity Generating Board (CEGB). The necessary plant specific information on
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capacities has been compiled from various sources, among them CEGB [3] and UNIPEDE [IS]. The inventory may be considered as valid from the late 1980s until about 1995 because neither significant commissionings nor decommissionings are to be expected for that period, the latter being due to the fact that much of the capacity presently existing was commissioned in the late 1960s or later and is assigned a lifetime of at least 40 years (Camsey [2]). The number of plants considered is 86, of which 55 are coal- or lignite-fired and 31 oil-fired. (Electricity production from other fuels does not generate significant SO, emissions.) Their total capacity is 57 GW which is in line with the overall capacity figure given in IEA [l 11. Annual electricity output is computed as the product of capacity and annual operating hours, h. The latter are derived from projections of aggregate capacity and output by fuel type, as documented in IEA [lo]. Hence, no plant specific differences can be accounted for. The figures are 5 700 hours for solid fuels and 500 hours for oil. They reflect the fact that coal stations are run on base and intermediate load, whereas oil-fired stations are limited to high demand periods. The resulting total electricity output is 234.8 TWh per year. Annual SO, emissions of the ith power plant are calculated using the above figures for operating hours, plant specific capacity figures, and SO, emission factors, according to the equation $=g.MWi.hi
(1)
where the superscript o characterizes ‘original’ (nonabatement) emissions. .s; is the SO, emission factor, specifying the amount of SO, emissions per MWh of electricity produced. It is obtained as &;=I$(1 -e-@$’ where the variables on the right hand side are fuel specific parameters. The last term in brackets gives the fuel input coefficient, ie the quantity of fuel necessary to produce one MWh. It is inversely related to the net calorific value cvi of the fuel and the power plant efficiency e, (3 600 is the conversion factor between MJ and MWh, which has to be applied because cvi is expressed in MJ). g denotes the sulphur content and 4 the sulphur retention factor of the fuel, cry is a fuel specific multiplier. The figures used to compute 4 are listed in Table 4 in the Appendix. They have been taken from CIAB [4] and Smith [17]. The resulting emission factors, expressed as kg/MWh, are 12.042 for coal, 11.382 for lignite and 16.645 for oil. Total annual SO, emissions of the British electricity sector are obtained by summing sy over all i. They
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October 1988
amount
to 2.865 x lo6 t for any of the years considered and are slightly lower than the 1980 figure of 3.02 x lo6 t reported in CIAB [4]. The difference can partly be explained by the fact that the latter figure also includes emissions of plants with a capacity of 5&100 MW. However, there are also reasons for an actual decrease of emissions until the early 1990s in particular the reduction in the projected input share of oil. In fact, the figure obtained is in good accordance with the forecast emissions of about 2.95 x lo6 t reported in Highton and Webb [7] for a scenario of low GDP growth.
Modelling abatement costs under alternative regulatory strategies As mentioned earlier, the dominant technology to achieve SO, abatement rates higher than about 20% is flue gas desulphurization (FGD). The specific FGD process most widely used is the limestone-gypsum process. In West Germany, eg where there has been a large FGD investment programme in recent years, this method is employed in more than 90% of all cases (Schlrer and Haug 1151). In fact, the limestonegypsum process is also the one used in Britain’s first FGD unit (see Power Europe, 18 June 1987, p 3). For these reasons, the following considerations and computations will be based on the limestone-gypsum FGD process, with an SO, removal rate of 95 %. However, it can be supposed that the comparative conclusions regarding regulatory strategies are not too sensitive with respect to the choice of the FGD process, because all processes are characterized by a large, elatively uniform amount of set-up costs (see United Nations [20] and IEA [12]), which is the crucial feature for the following analysis. The technical specification of the FGD cost function is based mainly on German experience, which is by now rather extensive. However, it was checked against British information if available. The principle source used is Scharer and Haug [ 153. Annual costs of the limestone-gypsum FGD process comprise annualized investment outlays, the cost of maintenance, repair, insurance and administration (considered to depend on investment outlays), labour costs, and the costs of water, electricity and limestone. Investment outlays, Z,depend on the (hourly) volume flow of flue gas to be desulphurized and on the (hourly) mass flow of SO,, denoted by Cand S, respectively. (In this section indices referring to individual plants are omitted for simplicity.) With respect to both determinants there is a decrease in unit investment, due to set-up costs. This property is captured by the linear specification: I=a+b,i,+b,S
(3)
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Annual labour requirements are considered to be a constant and denoted by f;, w denotes the unit labour cost. The annual electricity and water input is proportional to the annual flue gas volume v = #r. Electricity and water requirement per unit of v are denoted by rc,i and K, and the corresponding prices by P,, and P,. The annual limestone input is proportional to the annual SO, mass s = 3, lcls and P,, refer to the corresponding per unit input and price. Finally, letting 6 denote the annuity factor and y a status parameter that takes different values depending on whether the FGD unit is installed in a new plant or retrofitted to an existing one, total annual FGD costs can be written as:
c = 6(y(a+ bli, + b,S))+ wL+ (P,,K,~ + P,K,)irh
unit depends largely on the flue gas flow, it is less expensive to desulphurize a partial flow at maximum degree than the total flow to a lesser extent (Rentz [14]). Let 1 denote the fraction of the total flue gas flow to be subjected to FGD. i. is determined by the condition that the average of the SOZ concentrations in the desulphurized partial flow (( 1 - a)&‘) and in the by-pass flow (k”) equals K: 1(1-a)k”+(l-;l)k”=E This yields A= (1 - L/k”)/a
However, in order to assess the reaction of emitters to regulations relating to SO, abatement, costs have to be expressed in terms of the amount of SOZ abated. Denoting by a the maximum technically possible degree of abatement (ie the ratio of SO, removed to pre-abatement SO,) and by k” the pre-abatement concentration of SO, in flue gas, the annual amount of SO, removed is given by’
Thus, Equation
(5)
(4) can be rewritten in terms of r:
h
c = 6(y(a + ~ ak”h
I PIs% I a
(6)
Hence the cost function is linear in r with fixed costs of 6ya + wL, and the unit abatement cost decreases over the whole range up to the capacity limit P’” = a.?. The properties of the cost function determine the behaviour of the emitters under alternative types of regulation. Consider first a uniform limit on the SO, concentration in flue gas (concentration standard), denoted by 7i. The objective of any emitter is to minimize the cost of conforming to that standard. Owing to the fact that the size of the desulphurization
‘Note that FGD is always performed at the maximum degree of abatement, the choice variable being the flue gas volume flow actually subjected to FGD, see below.
290
(8)
Hence, the flue gas flow to be desulphurized is b = (( 1 - I;/k”)/a)ir”
r = f. h = aSh = ak”bh
(7)
(9)
where 9” denotes the total flow. The alternative regulatory tool to be considered here is transferable emission permits. Under such a policy emissions are only allowed to the extent that the emitter possesses permits, and the total number of permits is fixed. Each emitter has, the choice of either abating emissions or purchasing permits, and this choice will, of course, be governed by the individual abatement costs relative to the price of permits. In making this decision, set-up costs are also considered variable since they can be saved by the purchase of a sufficient quantity of permits. Taking into account that, given the cost function discussed above, the minimum of the unit abatement costs is at the capacity limit, the decision will be all-or-nothing. If unit abatement costs at the capacity limit are smaller than the price of a unit emission permit, maximum abatement will be chosen; otherwise, permits will be purchased for the total amount of emissions. Stricter emission limits, ie a smaller total amount of permits supplied, lead to an increase in the price of a unit emission permit. Hence, a continuous tightening of overall emission limits induces emitters to take up abatement measures according to the merit order of their unit abatement cost minimum. Minimum overall abatement costs are thus obtained, for each overall emission limit, by cumulating the costs of emitters according to that merit order. This provides the basic aggregation rule for the construction of a macroeconomic abatement cost function from linear individual abatement cost functions. In the next section, individual FGD costs will be determined numerically, and the implied aggregate cost function for SO, abatement in the British electricity sector will be derived.
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Numerical determination of SO, abatement costs
abatement
(f/EWb)
eb
Capacity NW)
ucJa VFg)
&Wb)
1152 1240 1344 1381 1808 1908 1915 1940 1948 2000 2304
0.278 0.277 0.276 0.275 0.272 0.272 0.272 0.271 0.271 0.271 0.270
3.168 3.155 3.150 3.114 3.107 3.107 3.105 3.105 3.102 3.088
eb
Capacity (MW)
&Wb)
plants
“UC’ denotes “UP denotes
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of capacity.
.b
Capacity (MW) Coal-fired 112 114 118 120 129 168 224 240 282 300 308 321 336 366 376 392 448 560 564 660 714 860 930 942 954 970 1100
costs as a function
control:
by means of the average 1985 exchange rate. Since all the stations included in the inventory have been commissioned no later than 1990, the resulting investment outlays are multiplied by a retrofit factor for which a value of 1.33 is assumed, in line with Schirer and Haug [15] and Smith [ 173. For the annuity factor, a value of 17 % was chosen, comprising an interest rate and a depreciation rate of 5%, respectively, 4 % for maintenance and repair, and 3 % for taxes, insurance and administration. The assumption on the interest rate is based on Smith [17], and the depreciation rate corresponds to a 20 years lifetime of FGD facilities, which is in accordance with the age structure of the power plants, as discussed earlier. Regarding labour costs, a personnel demand for two shifts (ie 4 000 hours per year) and hourly labour costs of E5 are assumed. The input coefficients of electricity, water and limestone, as listed in Table 5, are based on Schlrer and Haug [15]. It should be noted that the electricity requirement is rather low, due to the application of regenerative heat exchangers. In Table 1 unit costs of SO, abatement are presented as a function of capacity, grouped according to the type of fuel. The first column shows capacity in MW and the second one the costs of abating one unit of
According to Equation (4), the determinants of annual FGD costs are the (hourly) flows of flue gas C and SO, S to be subjected to desulphurization, and the annual operating hours. As discussed in the preceding section, b (and, accordingly, S) may be only fractions of the corresponding total flows, depending on the regulatory policy pursued. The total hourly flows, in turn, are computed as the product of capacity and fuel specific emission factors for flue gas and SO,, respectively (see Table 4 in the Appendix). The cost estimates presented in this paper are in pounds sterling as of 1985. They have been obtained by assigning numerical values to the input coefficients and prices of the variable factors and the parameters of the investment function as well as to the annuity factor and the status coefficient showing up in Equation (4). The figures are assembled in Table 5 in the Appendix. Methods and assumptions underlying their construction are briefly discussed. The coefficients of the investment function have been estimated by running a regression on German data and transforming the estimates into pounds sterling
Table 1. Unit SO,
policies for sulphur dioxide
0.430 0.427 0.42 1 0.419 0.408 0.374 0.346 0.340 0.329 0.324 0.323 0.320 0.318 0.313 0.312 0.310 0.304 0.295 0.295 0.290 0.288 0.284 0.282 0.282 0.281 0.28 1 0.279 unit abatement unit abatement
ECONOMICS
4.917 4.883 4.819 4.789 4.664 4.276 3.956 3.892 3.758 3.712 3.694 3.665 3.635 3.583 3.567 3.544 3.475 3.379 3.376 3.321 3.296 3.245 3.226 3.223 3.220 3.216 3.190
Lignite-fired 232 234 240
Oil-fired 102 110 124 140 202
3.181
plants 0.439 0.438 0.436
4.743 4.735 4.712
228 253’ 280 300 342 345 420 480 613 630 660 1080 1900 1920 1932
2.698 2.630 2.571 2.534 2.47 1 2.467 2.386 2.340 2.270 2.263 2.252 2.160 2.098 2.097 2.097
42.655 41.591 40.656 40.071 39.066 39.004 37.733 37.002 35.893 35.785 35.607 34.161 33.180 33.166 33.158
plants 3.539 3.428 3.269 3.126 2.785
55.957 54.206 51.686 49.424 44.041
costs relative to SO, removed. costs relative to electricity output.
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SO,, expressed in f/kg. For the quantity of SO, removed the maximum possible amount was assumed, ie 95 % of the original SO, quantity, according to the removal efficiency of the FGD process considered. The third column of the table contains the abatement costs per unit of output, expressed in f/MWh. It should be noted that the number of entries is smaller than the number of plants contained in the inventory because double entries of equal figures, stemming from identical plant and fuel characteristics were omitted. From the table it can be seen that both the unit costs relative to SO, removed (UC’)and the unit costs relative to electricity output (UP) differ substantially, depending on the fuel type. Within each fuel category UC’as well as uce increases as the capacity decreases. For coal-fired plants, UC’ranges from iO.27/kg for a plant of 2304 MW to f0.43 in the case of a 112 MW plant. The three lignite-fired plants have approximately the same capacity, requiring costs that range from f0.436/kg (240 MW) to fO.439/kg (232 MW). For oil-fired plants, UP is substantially higher in all cases, the minimum being f2.097/kg for a 1932 MW plant and the maximum &3.539/kg for 102 MW. To put these observations in perspective, it may be noted that, with the type of cost function given in Equation (6), UC”is always inversely related to capacity, operating hours and the SO, emission factor. (This can be shown by writing the amount of sulphur removed in terms of MW, h, sS and the removal rate, see Weizsacker and Welsch [21].) However, in our data, h, and sS are only distinguished by fuel type, not on a plant-by-plant basis. Moreover, a uniform, high value is assigned to h for both coal and lignite, whereas for oil h is lower by a factor of 10. On the other hand, E’for coal and lignite is only slightly different, whereas the corresponding value for oil is considerably higher, due to the higher sulphur content. These circumstances provide an explanation of the differences in ucs between fuel types, showing up in Table 1: The observation that UC”of the smallest coal-fired station is far smaller than UC’of the largest oil-fired plant reflects the fact that the difference in operating hours between solid fuels and oil is much larger than that in the corresponding emission factors and that the combined effect of these two differences outweighs economies related to capacity. The difference in UC”between coal and lignite fired plants, being caused only by a small difference in sS, is less pronounced. Regarding unit costs relative to electricity output, the picture is similar but not identical to the one just described. In particular, it is not universally true that coal-fired plants have lower UP than lignite-fired ones. However, for a given size (240 MW) UP is definitely lower in the case of coal firing. Regarding oil-fired
292
plants, it may be noted that FGD costs per MWh are in the same order of magnitude as average revenue of electricity delivered to industry. In order to assess the reliability of the figures obtained it might be desirable to check them against estimates from other sources. However, the information on FGD costs in the UK is rather scarce because, up to the middle of 1987, only one FGD unit had been installed in the UK electricity sector. The only useful source known to the author is Highton and Webb [7]. There, an estimate of 0.330 pounds per kg of SO, abated is reported for a power station to which FGD is retrofitted. According to Highton and Webb [S] this figure applies to a 2000 MW coal-fired plant with characteristics comparable to our assumptions. However, the estimate is based on an FGD process with only 80 % removal efficiency, in contrast to the 95 % assumed in the above calculations. Hence the amount of sulphur removed is smaller and economies of scale implied by set-up costs are less effective. Given this difference in the underlying assumptions and the technical progress that may have taken place since their estimate was made (eg the use of regenerative heat exchangers), their estimate seems to support our results. Of course, in view of the main purpose of this paper, the determination of SO, abatement costs on a plant-by-plant basis is only an intermediate step. As outlined earlier, the estimates of plant specific costs per unit of SO, removed serve as a basis for the construction of an aggregate cost function of SO, abatement in the British electricity sector. Table 2 shows, in the first two columns, the results of cumulating the quantities of SO, removed as well as the corresponding costs, according to the merit order of unit removal costs. The set of pairs of these two cumulated variables provides the aggregate SO, abatement cost function. Note the sharp cost increase in the range of about 2.59 x lo6 t of SO, to be removed. This marks the move in the merit order from solid-fired plants to oil-fired ones. In the marginal costs, shown in the third column, this jump is even more striking. The aggregate cost function provides a basis for the cost comparison of the two alternative regulatory strategies for SO, control to be carried out next.
Policy simulations The two types of regulation considered are uniform limits on the SO, concentration in flue gas (concentration standard) and transferable emission permits. The first is the regulatory tool most commonly employed in controlling SO, emissions (see OECD [ 131 for a compilation of concentration standards valid in various countries). The UK, however, has up till
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Table 2. Minimum
aggregate
abatement
costs as a function
of SO2 removed. Marginal
Marginal SO, removed (1 x 103) 1SO.239 276.742 403.768 534.183 664.599 782.495 906.912 1031.785 1121.838 1209.477 1290.335 1365.454 1437.183 1499.392 1562.643 1623.287 1684.713 1740.792 1787.350 1830.387 1873.424 19 16.462 1959.499 2002.536 2045.573 2082.090 2118.867 2118.080 2177.293 2206.507 2232.068
Absolute costs (f x 106) 40.552 74.892 109.368 144.731 180.095 212.182 245.975 279.888 304.686 328.853 351.245 372.135 392.137 409.648 427.43 1 444.532 461.838 477.744 491.158 503.650 516.143 528.635 541.128 553.620 566.112 576.898 587.752 596.626 605.500 614.374 622.292
COSk3
%%)
0.271 0.271 0.271 0.271 0.272 0.272 0.272 0.275 0.276 0.277 0.278 0.279 0.281 0.28 1 0.282 0.282 0.284 0.288 0.290 0.290 0.290 0.290 0.290 0.290 0.295 0.295 0.304 0.304 0.304 0.310
SO2 removed (t x 103) 2256.586 2281.105 2304.97 1 2326.881 2347.8 13 2367.896 2387.459 2405.848 2421.498 2436.104 2450.711 2465.318 2476.272 2487.227 2495.639 2503.464 2511.159 2518.592 2526.026 2533.459 2540.763 2555.555 2569.977 2584.276 2599.457 2614.732 2629.754 2638.293 2643.511 2648.729 2653.948
Absolute costs (f x 106) 629.937 637.582 645.057 652.019 658.726 665.210 671.558 677.599 682.923 687.973 693.024 698.075 702.170 706.265 709.694 712.970 716.212 719.385 722.558 725.73 1 728.871 735.316 741.631 747.903 779.743 811.774 843.294 861.741 873.492 885.242 896.993
now pursued a policy of dispersing emissions by means of tall stacks rather than abating them, implying that a large part of SO2 emitted is deposited outside its boundaries. Taking for granted that this necessitates a certain overall reduction in UK emissions, the question arises how such overall reduction targets are best being implemented. The variability of FGD costs per unit of SO, removed, as discussed in the previous section, indicates a large potential for saving costs by employing cost-effective regulatory strategies instead of uniform emission limits. In what follows, the extent to which advantage can be taken of this potential will be examined. The method used is to compute the costs ofachieving various overall reduction targets by means of either concentration standards or transferable emission permits. The simulations presented are partly based on reduction plans currently aired. One is the agreement on a 30% reduction relative to 1980 emissions mentioned in the introduction. Another is the 60 % reduction, also based on 1980 emissions, according to the proposed EC Directive on the limitation of emissions from large
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CO.@ wk)
0.312 0.312 0.313 0.318 0.320 0.323 0.324 0.329 0.340 0.346 0.346 0.346 0.374 0.374 0.408 0.419 0.421 0.427 0.427 0.427 0.430 0.436 0.438 0.439 2.097 2.097 2.098 2.160 2.252 2.252 2.252
SO, removed (t x 103) 2659.166 2664.384 2669.602 2674.821 2679.802 2684.648 2689.495 2693.290 2696.611 2699.338 2702.042 2704.414 2706.786 2709.000 2711.000 2712.803 2714.400 2715.507 2716.614 2717.721 2718.828 2719.808 2720.678 2721.484
Absolute costs (f x 106) 908.743 920.494 932.244 943.995 955.267 966.268 977.269 986.150 994.074 1000.802 1007.482 1013.493 1019.503 1025.195 1030.457 1035.319 1039.768 1043.227 1046.687 1050.147 1053.606 1056.811 1059.792 1062.646
Margioal costs (wz) 2.252 2.252 2.252 2.252 2.263 2.270 2.270 2.340 2.386 2.467 2.470 2.534 2.534 2.571 2.630 2.697 2.785 3.125 3.125 3.125 3.125 3.268 3.428 3.539
combustion plants. For each of these reduction plans, that uniform concentration limit which is necessary to achieve the given reduction target is derived. Note that the EC Directive, in addition to the 60% reduction rate, contains explicit concentration limits, namely 400 mg/m3 to be applied as from 1 January 1985 and 250mg/m3 applying after 31 December 1995. In the case of the UK, these are stricter than necessary for a 60% emission reduction. Therefore, a second set of simulations is run, based on the reductions resulting from these concentration limits. Finally, in order to provide a sensitivity analysis, various notional standards, specifying reduction rates relative to projected annual emissions in the reference period, are considered. The results concerning costs were obtained by modelling the emitters’ reaction to regulation in the way described earlier. This means that in the case of concentration standards the fraction of flue gas actually to be subjected to desulphurization is endogenously determined as a function of pre-abatement concentration, removal efficiency and concentration standard,
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H.
Welsch
according to Equation (8). Overall abatement costs are then obtained by adding the abatement costs of all power stations. In the case of transferable emission permits it is assumed that the relatively most efficient plants (in terms of unit SO, removal costs) desulphurize their complete flue gas flow and the others take no abatement measures at all, the two subsets of plants being separated by the condition that aggregate SO, abatement conforms to the given reduction target. Overall abatement costs are then obtained according to the aggregate cost function tabulated in Table 2. The scenarios considered and the simulation results are summarized in Table 3. The leftmost part of the table contains the specification of the scenarios in terms of concentration standards iG,emission reduction rates p, the amount of SO, removed, R, and the quantity of post-abatement emissions, S. Beginning with the sixth column, the costs of achieving the removal targets R by means of either uniform concentration standards (CS) or transferable emission permits (TEP) are tabulated. C denotes total costs, UC” unit costs relative to SO2 removed and UC’ unit costs relative to electricity output. TCS, reported in the last but one column, is total costs saved by TEP, and CR is the percentage ratio of total costs under TEP and under cs. The first two scenarios are defined by concentration standards of 400 mg/m3 and 250 mg/m3 according to the EC Directive. The implied quantities of SO, to be removed are R = 2.548 x 106t and R = 2.667 x lo?, respectively. Relating these to the non-abatement emissions of 2.865 x 106t derived earlier gives abatement rates of p = 88.9 % and p = 93.1%. The postabatement emissions are S = 0.317 x 106t and S= 0.198 x 106t. Scenarios 3 and 4 refer to 30 % and 60 %
reductions relative to 1980 emissions. Since the latter amounts to 3.020 x lo’?, the implied reduction rates relative to the non-abatement emissions in the reference period are somewhat smaller, namely p = 26.2 and p = 57.8. The concentration standards that would be necessary to achieve these reductions are E = 2667 mg/m3 and 7t = 1524 mg/m3, respectively. Scenarios 5-11 are based on notional percentage reduction targets ranging from 30% to 90%. Turning to the results, the prime observation to be made is that there are in fact large cost savings to be achieved by TEP. Referring to reduction rates from 3&80%, total costs saved increase from 2208.795 to f279.890 million, while beyond 80% TCS decreases (see scenarios 1, 2 and 11). For p = 93.1% they are f 115.729 million, and it is evident that for the maximum overall reduction rate of 95% (as determined by the FGD process specification) costs under TEP are equal to those under CS, because in this case the allocation of abatement activities is identical under both regimes. Regarding the relative cost advantage, it can be seen that the cost ratio increases as reduction targets become more demanding, approaching, of course, 100 % as the reduction ratio approaches its maximum. Note that this increase is relatively flat over much of the range and becomes steeper only for reduction rates higher than about 90%. However, considering the proposed EC Directive these high reduction rates are politically very relevant. From the table it can be seen that overall reductions equivalent to the uniform concentration standard of 400 mg/m3 required in the Directive for 1985-95 can be achieved under TEP at costs that are 27.1% lower than under CS (scenario 1). In terms of total cost savings, the figure of f273.63 1
Table 3. Overall SO2 abatement costs resulting from policy simulations. Specification of Scenarios No I; P (m&m? W) 1 400 88.9 2 250 93.1 3 2661 26.2’ 4 1524 57.Sb 5 2530 30.0 6 2165 40.0 7 1807 50.0 8 1446 60.0 9 1084 70.0 10 123 80.0 11 361 90.0 Note: For notations ‘Corresponds Vorresponds
294
Costs under CS
2667 751 1657 860 1146 1433 1719 2005 2292 2578
198 2114 1208 2005 1719 1433 1146 860 513 287
1047.974 419.090 716.487 454.770 548.805 642.841 736.876 830.911 924.947 1018.983
0.393 0.558 0.432 0.529 0.479 0.449 0.429 0.414 0.404 0.395
Costs under TEP UC’ (f/MWb) 4.293 4.463 1.785 3.052 1.937 2.337 2.738 3.138 3.529 3.939 4.343
Coat comparison
F3;.;p,^’
;;!I
932.244 212.182 461.838 245.975 328.853 392.137 477.744 566.112 645.057 741.903
0.349 0.271 0.214 0.271 0.272 0.273 0.274 0.277 0.280 0.289
$kWb) 3.132 3.970 0.904 1.967 1.084 1.401 1.670 2.035 2.411 2.141 3.185
:“x 106) 273.63 1 115.729 206.908 254.650 208.795 219.952 250.703 259.132 264.799 279.890 271.080
:‘, 72.9 89.0 50.6 64.5 54.1 59.9 61.0 64.8 68.1 69.7 73.4
see text.
to a 30% reduction relative to 1980 emissions. to a 60% reduction relative to 1980 emissions.
ENERGY
ECONOMICS
October 1988
Alternative
million is close to the maximum of TCS attainable under all scenarios. Reductions resulting from a uniform standard of 250mg/m3, as envisaged for the time after 1995, would still allow for cost savings of 11 %, totalling 0 15.729 million (scenario 2). The differences in total costs between the two types of regulation are of course reflected in unit costs. However, it should be noted that the figures for UC under TEP are not exactly the product of their counterparts under CS and the ratio of total costs (CR). This is due to the fact that UC” under TEP is based on total SO2 reductions which are generally higher than those under CS. The reason for this is that, for simplicity, it was assumed that the marginal plant taking abatement measures under TEP performs desulphurization at full degree. It is not generally clear whether the figures given for UC under TEP overestimate or underestimate those valid for exactly matching the overall reduction targets. However, with respect to absolute figures, the above simplification implies an overestimation of total costs to be spent for reaching a given target under TEP. On the other hand, costs under CS are probably underestimated because a precise calibration of the fraction of flue gas to be subjected to desulphurization has been assumed. In practice, it is very likely that the desulphurization unit will be more or less oversized for reasons of uncertainty. Thus, the estimates of total cost savings seem to be on the safe side. It remains to comment briefly on desulphurization costs per unit of electricity produced. For high reduction rates (7&90%) the order of magnitude of UC” under CS is roughly one tenth of the price of electricity. Under TEP it could be about 30% lower, ie 7% of the electricity price. For 30% abatement the price increase induced by FGD is about 5% under CS, while it could be as low as 2.5%.
Conclusions This paper has been concerned with the costs of achieving overall reduction targets for sulphur dioxide emissions of the British electricity sector under two alternative regulatory strategies: uniform limits to SOZ concentrations in flue gas and transferable emission permits. The motivation for this investigation lies in the fact that the UK, being the largest SO, producer in Western Europe, would be particularly affected by political initiatives currently taken to reduce SO* emissions all over Europe. One such initiative is the proposal for an EC Directive on the limitation of emissions from large combustion plants. It contains the objective of reducing SO2 emissions by 60% relative to the 1980 figures and specifies a concentration standard of 400mg/m3 applying from 1985-95. For
ENERGY ECONOMICS
October 1988
policies jar sulphur dioside
control:
H. WeDch
the UK, this concentration limit would imply an emission reduction rate of 88.9% relative to annual emissions in the early 1990s. The question examined in this paper is, to what extent such reduction rates could be achieved in a more cost-effective way by employing the incentive mechanism implied by the polluter-pays principle. Note, in this context, that in the UK local effects of SO, emissions are alleviated by means of tall stacks. In order to control distant effects, on the other hand, it is sufficient to concentrate on overall emissions. The analysis is directed to the power plant sector mainly because it has a major share in the SO, emissions in the UK and because the concentation limits mentioned above explicitly refer to it. Regarding the results, the first statement to be made refers to SO, abatement technology. The estimates obtained indicate that, in recent years, improvements in flue gas desulphurization technology (which is the only one relevant to achieve reduction rates in the range considered) seem to have induced a certain decrease in the unit cost of SO, removal to be borne by individual plants. Second, relating to regulatory strategies, the properties of the FGD cost function imply that under transferable emission permits a given plant either performs desulphurization on maximum scale or takes no abatement measures at all. The cost savings obtainable from such a reallocation of abatement activities relative to a regime of concentration limits are in the range of 5&30% for emission reduction rates of 30-80%. Thus, percentage cost savings are inversely related to reductions in emissions. In absolute terms, however, cost savings increase from E209 x lo6 to ;E280 x lo6 as reduction rates increase from 3(X80%. Beyond 80%, not only percentage but also absolute costs saved decrease. Nevertheless, they are considerable. For example, the 88.9% reduction implied by the concentration limit of 400mg/m3 contained in the EC Directive could be achieved, under transferable emission permits, at costs that are 27.1% or almost E274 x lo6 lower. Of course, these estimates are only rough projections of the most likely outcomes of the policies considered. This is mainly due to the fact that differences in the efficiency, operating hours and characteristics of fuels burnt by individual plants could only incompletely be accounted for, due to lack of data. However, since the cost saving potential of transferable emission permits relative to uniform emission limits is based precisely on such differences, the estimation bias is probably on the safe side. Thus, it may be concluded that costs to be saved by the British electricity sector are considerable. Whether this justifies a step away from traditional regulatory patterns can, of course, not be decided on these grounds alone.
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The results concerning aggregate abatement costs are not only relevant for the decision on regulatory strategies to implement overall reduction targets but also for setting targets in the context of a coordinated SO, abatement policy on a European scale. Note in this context that the aggregate cost function derived is characterized by increasing marginal costs over the whole range, showing an upward jump at that removal quantity for which the desulphurization of oil-fired plants becomes necessary. The latter would be the case for the overall abatement implied by a concentration limit of 250mg/m3, as envisaged in the EC Directive for the time after 1995, but not for the reduction associated with a limit of 400mg/m3, being relevant for the early 1990s. As a consequence, the corresponding increase in the quantity removed of less than five percentage points would require an increase in minimum costs of more than 20%. This seems worth taking into account when setting emission targets for countries participating in an internationally coordinated SO, control policy. Alternatively, it could provide an incentive to smooth the cost function by changing the fuel mix. In either case, transferable emission permits should not be excluded from consideration when thinking about how to reduce the effects of British SO, emissions in a cost-effective way.
6
7
8
9
10 11 12 13 14 15
16
17
18
References S. E. Atkinson and D. H. Lewis, ‘A cost-effectiveness analysis of alternative air quality control strategies’, Journal of Environmental Economics and Management, Vol 1, 1974, pp 237-250. G. T. B. Camsey ‘Strategie der CEGB zur ErhGhung der Lebensdauer von Kraftwerken’, VGB Kraftwerkstechnik, Vol 64, 1984, pp l-5. CEGB (Central Electricity Generating Board), Statistical Yearbook 1984185. CIAB (Coal Industry Advisory Board), Environmental Strategies for Coal Use, Paris, 1985. EC (European Community), ‘Proposal for a Council Directive on the limitation of emissions of pollutants into the air from large combustion plants, Oficial Journal of the European Communities No C 49/l, 1984.
19
20 21
22
N. H. Highton and M. G. Webb, ‘Sulphur dioxide from electricity generation: policy options for pollution control’, Energy Policy, Vol 8, 1980, pp 61-76. N. H. Highton and M. G. Webb, ‘Pollution abatement costs in the electricity supply industry in England and Wales’, Journal of Industrial Economics, Vol 30, 1981. pp 49-65. N. H. Highton and M. G. Webb, ‘The effects on electricity prices in England and Wales of notional sulphur dioxide emission standards for power stations’, Journal qf Environmental Economics and Management, Vol 11, 1984, pp 7@83. L. Hordijk, ‘Towards a targetted emission reduction in Europe’, Atmospheric Environment, Vol 20, 1986, pp 2053-2058. IEA (International Energy Agency), Electricity in ZEA Countries, Paris, 1985. IEA, Coal Information, Paris, 1986. IEA, FGD-System Performance, Paris, 1986. OECD, Emission Standards ,f;)r Major Air Pollutants, Paris, 1984. Otto Rentz, Techno-okonomie betriehlicher Emissionsminderungsmafinahmen, Berlin, 1979. B. SchHrer and N. Haug ‘The cost of flue gas desulphurization and denitrification in the Federal Republic of Germany’, Economic Bulletin,for Europe, Vol 39, 1987. E. P. Seskin, R. J. Anderson and R. 0. Reid, ‘An empirical analysis of economic strategies for controlling air pollution’, Journal of Environmental Economics and Management, Vol 10, 1983, pp 112-124. T. F. Smith, ‘Factors affecting costs of power station emission controls with particular reference to the UK? in OECD, ed, Energy and Cleaner Air, Paris, 1986. UNIPEDE (Union Internationale des Producteurs et Distributeurs d’Energie Electrique), Programmes and Prospects for the Electricity Sector 1984-1990 and 1990-1995. 1985. United Nations, ‘EMEP: The co-operative programme for monitoring and evaluation of long-range transport of air pollutants in Europe’, Economic Bulletin for Europe, Vol 34, 1982, pp 2940. United Nations, Air Pollution Across Boundaries, New York, 1985. C. C. von Weizsgcker and H. Welsch, ‘Alternative Umweltstrategen fiir die europlische Elektrizitltswirtschaft: Zur Kontrolle grenziiber schreitender externer Effekte von SO,-Emissionen, in H. Siebert, ed, Um~veltschut-_,fir I,@ und Wasser, Berlin, Heidelberg, New York, 1988, pp 97-130. H. Welsch, Regulatory Strategies to Control Transhoundarx Sulphur Dioxide Pollution, forthcoming.
Appendix Table 4. Basic data for estimation
of SO, emissions
r~’sulphur content [X/100] cu (net) calorific value (MJ/‘kg] d sulphur retention factor [%/lOO] urn multiplier [kg/kg] e power plant efficiency [%/lOO] h operating hours [h/a] E*SO, emission factor [kg/M Wh] E”flue gas emission factor [m”/MWh]
296
Coal 0.016
23.4 0.05 1.9 0.369 5700 12.042 3363.391
Lignite 0.01
8.4 0.30 1.4 0.369 5700 11.382 4083.851
ENERGY ECONOMICS
Oil 0.029
42.25 0.00 2.0 0.287 500 16.645 3363.676
October 1988
Ahrnutiw
policirs~for
sulphur
dio.rid(> control:
H. l+ih-l~
Table 5. Basic data for estimation of FGD costs (Cf Equation (4)). annuity factor retrofit factor regression
coefficient
hourly labour costs annual labour requirements price of electricity unit input of electricity price of water unit input of water price of limestone unit input of limestone exchange rate Prices and exchange
ENERGY
0.17 1.33 f5340655.22 f8.80858 f2349.6775 f5 4000 hours f36/MWh 0.31 10-s MWh/ms flue gas 1 DM/ms 0.369 lo-“ m3/m3 flue gas fO.O152/kg 1.56 kg/kg SO, f0.278/DM
rates refer to 1985.
ECONOMICS
October
1988
297
Heinz Welsch
The paper is concerned with the costs of achieving overall reduction targets for sulphur dioxide (SO,) emissions of the British electricity sector under two alternative regulatory strategies: unifarm limits to SO, concentration in$ue gas and transferable emission permits. The implications of the SO, abatement technology for the allocation of abatement activities under the two types of regulation are analysed, and overall abatement costs are compared. An aggregate cost function is developed which may serve, together with transboundary SO2 transfer coeficients, as a basis for eficiently setting overall emission reduction targets for countries participating in a coordinated SO, control policy. Keywords: Airquality management; Emission abatement costs; cost-effectiveness analysis
In the past one and a half decades the problem of transboundary air pollution has received increasing attention. As a result of this, in 1977 the UN Economic Commission for Europe established the ‘Cooperative Programme for the Monitoring and Evaluation of Long-Range Transmission of Air Pollutants in Europe’ (EMEP). From the atmospheric model developed within EMEP (see United Nations Cl93 a matrix of transfer coefficients can be derived which relates depositions of sulphur dioxide (SO,) in 27 European countries to SO, emissions in any one of these countries. This matrix reveals that for most of the countries more than half of the SO* emissions are exported. Regarding depositions, in most cases more than half are of foreign origin. For some countries, in
The author is with the Institute of Energy Economics, University of Cologne, Albertus-Magnus-Platz, D-5000 K6ln 41, FR Germany. Numerical results presented in this paper are based on data partly compiled by I. Hoven, W. Schulz and P. H. Suding. Their contribution is gratefully acknowledged, while the sole responsibility for any errors remains with the author. Final manuscript
received 15 April 1988.
0140-9883/88/040287-11
$03.00 0
1988 Butterworth
particular the Scandinavian ones, the ‘import share’ is considerably higher. Due to these findings the view was widely accepted that SO, control strategies have to be internationally coordinated. In 1984 a number of European countries and Canada agreed on reducing their SO, emissions by a uniform rate of 30 % in 1993 based on 1980 figures. In 1983 the Commission of the European Communities submitted a proposal for a Council Directive (EC [S]) requiring that Member States reduce annual SO, emissions from large combustion plants by 60% in 1995 relative to their 1980 emissions. Although, from a political point of view, these agreements are an achievement, one may wonder whether such uniform rates of emission reduction are an efficient way of reducing the effects of sulphur deposition. Using depositions in certain target areas as an indicator of damage, Hordijk [9] shows that the same degree of damage reduction which results from flat rate emission abatement all over Europe can also be achieved by selectively lowering emissions of certain countries, requiring a smaller decrease in overall emissions. This provides evidence that neglecting differences in the marginal damage caused by emissions of different origins, as expressed by unit transfer coefficients, may give rise to inefficiencies. However, from an economic point of view, this is only one side of
& Co (Publishers)
Ltd
287
Alternative
policies,fi)r
sulphur dioxide
control:
H. Welsch
the coin, since for economic optimization differences in marginal damage have to be balanced against differences in the marginal costs of emission abatement. Hence, to set efficiently emission reduction rates for the individual countries involved in a coordinated deposition-oriented SO, control policy it is not only necessary to know the matrix of transfer coefficients but also the ‘macroeconomic’ SO, abatement cost functions of these countries. Moreover, to guarantee efficiency, these target rates must then be implemented in a cost-effective way. Regarding this issue, it is well-known that the traditional regulatory strategies, such as prescribing uniform emission limits for individual emitters, are not cost-effective. This paper relates to both the efficient setting and implementation of emission reduction targets; however, it focuses on the latter. Its main purpose is to provide a nationwide cost comparison of two regulatory strategies for SO, control: uniform limits to the concentration of SO, pollutants in flue gas and transferable emission permits. Such a comparison is important because, though the superiority of transferable permits is theoretically evident, it is not a priori clear whether the amount of potential cost savings would justify a move away from traditional regulatory practice. Moreover, since transferable permits lead to a least-cost allocation of abatement activities for any given overall abatement level, a nationwide cost function of SO, emission abatement is obtained as a by-product of analysing the costs of a transferable emission permit policy. As mentioned above, such a cost function, if available for all countries participating in a coordinated SO, control policy, could serve, together with the transfer matrix, as a basis for efficiently setting emission abatement rates for the countries, given prespecitied deposition reduction targets. In the present paper the analysis is only carried out for one country as a leading example, namely the UK. For a more comprehensive analysis which covers all countries of Western Europe and their atmospheric linkages see Welsch [22]. Moreover, the scope of this study is restricted to the power plant sector. The reasons for this choice are as follows. According to the results obtained within EMEP, the UK is the largest producer of SO, in western Europe, emitting a total of 4.25 x lo6 t in 1982. Due to the policy currently pursued in the UK of alleviating local effects of SO, emissions by means of tall stacksand the prevalence of south-westerly winds, 63.4% of the total emissions or 2.7 x 1Oht are deposited outside its boundaries, mostly in continental Europe. Thus, in order to reduce depositions on the continent, the UK is a prime canditate for substantial reduction of its SO, emis-
288
sions, and an anaklysis of the costs of emission abatement in the UK is particularly called for, taking into consideration alternative regulatory strategies for implementing overall emission reduction targets. Since the cost comparison of such policies relies on modelling the behaviour of individual emitters, it is necessary to keep their number in a manageable range and to make sure that data on individual emitters are available. Both requirements are fulfilled by the power plant sector. Taking into account that in 1982 it accounted for 65.6% of SO, emissions in the UK, it seems justifiable to confine oneself to that industry. This also has the advantage that there exist several previous studies on SO, abatement options in the British electricity sector, which may serve as a basis to build upon and, partly, as a standard for checking results. In particular, the work of Highton and Webb [6,7,8] has to be mentioned. However, their research focuses on the cost-effective choice of abatement technologies rather than on the assessment of regulatory strategies, and the overall reduction rates they consider are only in the range of up to 40% by 1995. Nevertheless, their results are valuable for the present investigation because, a priori, one might think that the type of regulation influences the choice of technology and the latter has to be accounted for in a cost evaluation of the former. But, following Highton and Webb [7,8], the abatement potential of technologies such as coal or oil desulphurization is rather limited, and flue gas desulphurization (FGD) seems to be the only suitable method to achieve reduction rates higher than about 20%. Therefore, given the reduction rates currently envisaged, the following cost comparison of regulatory strategies will be based on one technology only, namely FGD. It may be noted that cost comparisons of the kind carried out in this paper are not a new type of undertaking. There have been numerous such studies relating to the US bubble policy. A classic example is Atkinson and Lewis [ 11, and a more recent one Seskin, Anderson and Reid [ 161. However, to the author’s knowledge, no such analysis has been performed on a nationwide scale, as is necessary for any attempt at an efficient, coordinated control of transboundary air pollution.
Sulphur dioxide emissions in the British electricity sector The projections of annual SO, emissions refer to the early 1990s. They are based on a power plant inventory comprising coal- and oil-fired plants with a capacity of 1OOMW or more, the majority being operated by the Central Electricity Generating Board (CEGB). The necessary plant specific information on
ENERGY ECONOMICS
October 1988
capacities has been compiled from various sources, among them CEGB [3] and UNIPEDE [IS]. The inventory may be considered as valid from the late 1980s until about 1995 because neither significant commissionings nor decommissionings are to be expected for that period, the latter being due to the fact that much of the capacity presently existing was commissioned in the late 1960s or later and is assigned a lifetime of at least 40 years (Camsey [2]). The number of plants considered is 86, of which 55 are coal- or lignite-fired and 31 oil-fired. (Electricity production from other fuels does not generate significant SO, emissions.) Their total capacity is 57 GW which is in line with the overall capacity figure given in IEA [l 11. Annual electricity output is computed as the product of capacity and annual operating hours, h. The latter are derived from projections of aggregate capacity and output by fuel type, as documented in IEA [lo]. Hence, no plant specific differences can be accounted for. The figures are 5 700 hours for solid fuels and 500 hours for oil. They reflect the fact that coal stations are run on base and intermediate load, whereas oil-fired stations are limited to high demand periods. The resulting total electricity output is 234.8 TWh per year. Annual SO, emissions of the ith power plant are calculated using the above figures for operating hours, plant specific capacity figures, and SO, emission factors, according to the equation $=g.MWi.hi
(1)
where the superscript o characterizes ‘original’ (nonabatement) emissions. .s; is the SO, emission factor, specifying the amount of SO, emissions per MWh of electricity produced. It is obtained as &;=I$(1 -e-@$’ where the variables on the right hand side are fuel specific parameters. The last term in brackets gives the fuel input coefficient, ie the quantity of fuel necessary to produce one MWh. It is inversely related to the net calorific value cvi of the fuel and the power plant efficiency e, (3 600 is the conversion factor between MJ and MWh, which has to be applied because cvi is expressed in MJ). g denotes the sulphur content and 4 the sulphur retention factor of the fuel, cry is a fuel specific multiplier. The figures used to compute 4 are listed in Table 4 in the Appendix. They have been taken from CIAB [4] and Smith [17]. The resulting emission factors, expressed as kg/MWh, are 12.042 for coal, 11.382 for lignite and 16.645 for oil. Total annual SO, emissions of the British electricity sector are obtained by summing sy over all i. They
ENERGY ECONOMICS
October 1988
amount
to 2.865 x lo6 t for any of the years considered and are slightly lower than the 1980 figure of 3.02 x lo6 t reported in CIAB [4]. The difference can partly be explained by the fact that the latter figure also includes emissions of plants with a capacity of 5&100 MW. However, there are also reasons for an actual decrease of emissions until the early 1990s in particular the reduction in the projected input share of oil. In fact, the figure obtained is in good accordance with the forecast emissions of about 2.95 x lo6 t reported in Highton and Webb [7] for a scenario of low GDP growth.
Modelling abatement costs under alternative regulatory strategies As mentioned earlier, the dominant technology to achieve SO, abatement rates higher than about 20% is flue gas desulphurization (FGD). The specific FGD process most widely used is the limestone-gypsum process. In West Germany, eg where there has been a large FGD investment programme in recent years, this method is employed in more than 90% of all cases (Schlrer and Haug 1151). In fact, the limestonegypsum process is also the one used in Britain’s first FGD unit (see Power Europe, 18 June 1987, p 3). For these reasons, the following considerations and computations will be based on the limestone-gypsum FGD process, with an SO, removal rate of 95 %. However, it can be supposed that the comparative conclusions regarding regulatory strategies are not too sensitive with respect to the choice of the FGD process, because all processes are characterized by a large, elatively uniform amount of set-up costs (see United Nations [20] and IEA [12]), which is the crucial feature for the following analysis. The technical specification of the FGD cost function is based mainly on German experience, which is by now rather extensive. However, it was checked against British information if available. The principle source used is Scharer and Haug [ 153. Annual costs of the limestone-gypsum FGD process comprise annualized investment outlays, the cost of maintenance, repair, insurance and administration (considered to depend on investment outlays), labour costs, and the costs of water, electricity and limestone. Investment outlays, Z,depend on the (hourly) volume flow of flue gas to be desulphurized and on the (hourly) mass flow of SO,, denoted by Cand S, respectively. (In this section indices referring to individual plants are omitted for simplicity.) With respect to both determinants there is a decrease in unit investment, due to set-up costs. This property is captured by the linear specification: I=a+b,i,+b,S
(3)
289
Alternative policies,for
sulphur dioxide control: H. We&h
Annual labour requirements are considered to be a constant and denoted by f;, w denotes the unit labour cost. The annual electricity and water input is proportional to the annual flue gas volume v = #r. Electricity and water requirement per unit of v are denoted by rc,i and K, and the corresponding prices by P,, and P,. The annual limestone input is proportional to the annual SO, mass s = 3, lcls and P,, refer to the corresponding per unit input and price. Finally, letting 6 denote the annuity factor and y a status parameter that takes different values depending on whether the FGD unit is installed in a new plant or retrofitted to an existing one, total annual FGD costs can be written as:
c = 6(y(a+ bli, + b,S))+ wL+ (P,,K,~ + P,K,)irh
unit depends largely on the flue gas flow, it is less expensive to desulphurize a partial flow at maximum degree than the total flow to a lesser extent (Rentz [14]). Let 1 denote the fraction of the total flue gas flow to be subjected to FGD. i. is determined by the condition that the average of the SOZ concentrations in the desulphurized partial flow (( 1 - a)&‘) and in the by-pass flow (k”) equals K: 1(1-a)k”+(l-;l)k”=E This yields A= (1 - L/k”)/a
However, in order to assess the reaction of emitters to regulations relating to SO, abatement, costs have to be expressed in terms of the amount of SOZ abated. Denoting by a the maximum technically possible degree of abatement (ie the ratio of SO, removed to pre-abatement SO,) and by k” the pre-abatement concentration of SO, in flue gas, the annual amount of SO, removed is given by’
Thus, Equation
(5)
(4) can be rewritten in terms of r:
h
c = 6(y(a + ~ ak”h
I PIs% I a
(6)
Hence the cost function is linear in r with fixed costs of 6ya + wL, and the unit abatement cost decreases over the whole range up to the capacity limit P’” = a.?. The properties of the cost function determine the behaviour of the emitters under alternative types of regulation. Consider first a uniform limit on the SO, concentration in flue gas (concentration standard), denoted by 7i. The objective of any emitter is to minimize the cost of conforming to that standard. Owing to the fact that the size of the desulphurization
‘Note that FGD is always performed at the maximum degree of abatement, the choice variable being the flue gas volume flow actually subjected to FGD, see below.
290
(8)
Hence, the flue gas flow to be desulphurized is b = (( 1 - I;/k”)/a)ir”
r = f. h = aSh = ak”bh
(7)
(9)
where 9” denotes the total flow. The alternative regulatory tool to be considered here is transferable emission permits. Under such a policy emissions are only allowed to the extent that the emitter possesses permits, and the total number of permits is fixed. Each emitter has, the choice of either abating emissions or purchasing permits, and this choice will, of course, be governed by the individual abatement costs relative to the price of permits. In making this decision, set-up costs are also considered variable since they can be saved by the purchase of a sufficient quantity of permits. Taking into account that, given the cost function discussed above, the minimum of the unit abatement costs is at the capacity limit, the decision will be all-or-nothing. If unit abatement costs at the capacity limit are smaller than the price of a unit emission permit, maximum abatement will be chosen; otherwise, permits will be purchased for the total amount of emissions. Stricter emission limits, ie a smaller total amount of permits supplied, lead to an increase in the price of a unit emission permit. Hence, a continuous tightening of overall emission limits induces emitters to take up abatement measures according to the merit order of their unit abatement cost minimum. Minimum overall abatement costs are thus obtained, for each overall emission limit, by cumulating the costs of emitters according to that merit order. This provides the basic aggregation rule for the construction of a macroeconomic abatement cost function from linear individual abatement cost functions. In the next section, individual FGD costs will be determined numerically, and the implied aggregate cost function for SO, abatement in the British electricity sector will be derived.
ENERGY ECONOMICS
October 1988
Alternative
Numerical determination of SO, abatement costs
abatement
(f/EWb)
eb
Capacity NW)
ucJa VFg)
&Wb)
1152 1240 1344 1381 1808 1908 1915 1940 1948 2000 2304
0.278 0.277 0.276 0.275 0.272 0.272 0.272 0.271 0.271 0.271 0.270
3.168 3.155 3.150 3.114 3.107 3.107 3.105 3.105 3.102 3.088
eb
Capacity (MW)
&Wb)
plants
“UC’ denotes “UP denotes
ENERGY
H. W&ch
of capacity.
.b
Capacity (MW) Coal-fired 112 114 118 120 129 168 224 240 282 300 308 321 336 366 376 392 448 560 564 660 714 860 930 942 954 970 1100
costs as a function
control:
by means of the average 1985 exchange rate. Since all the stations included in the inventory have been commissioned no later than 1990, the resulting investment outlays are multiplied by a retrofit factor for which a value of 1.33 is assumed, in line with Schirer and Haug [15] and Smith [ 173. For the annuity factor, a value of 17 % was chosen, comprising an interest rate and a depreciation rate of 5%, respectively, 4 % for maintenance and repair, and 3 % for taxes, insurance and administration. The assumption on the interest rate is based on Smith [17], and the depreciation rate corresponds to a 20 years lifetime of FGD facilities, which is in accordance with the age structure of the power plants, as discussed earlier. Regarding labour costs, a personnel demand for two shifts (ie 4 000 hours per year) and hourly labour costs of E5 are assumed. The input coefficients of electricity, water and limestone, as listed in Table 5, are based on Schlrer and Haug [15]. It should be noted that the electricity requirement is rather low, due to the application of regenerative heat exchangers. In Table 1 unit costs of SO, abatement are presented as a function of capacity, grouped according to the type of fuel. The first column shows capacity in MW and the second one the costs of abating one unit of
According to Equation (4), the determinants of annual FGD costs are the (hourly) flows of flue gas C and SO, S to be subjected to desulphurization, and the annual operating hours. As discussed in the preceding section, b (and, accordingly, S) may be only fractions of the corresponding total flows, depending on the regulatory policy pursued. The total hourly flows, in turn, are computed as the product of capacity and fuel specific emission factors for flue gas and SO,, respectively (see Table 4 in the Appendix). The cost estimates presented in this paper are in pounds sterling as of 1985. They have been obtained by assigning numerical values to the input coefficients and prices of the variable factors and the parameters of the investment function as well as to the annuity factor and the status coefficient showing up in Equation (4). The figures are assembled in Table 5 in the Appendix. Methods and assumptions underlying their construction are briefly discussed. The coefficients of the investment function have been estimated by running a regression on German data and transforming the estimates into pounds sterling
Table 1. Unit SO,
policies for sulphur dioxide
0.430 0.427 0.42 1 0.419 0.408 0.374 0.346 0.340 0.329 0.324 0.323 0.320 0.318 0.313 0.312 0.310 0.304 0.295 0.295 0.290 0.288 0.284 0.282 0.282 0.281 0.28 1 0.279 unit abatement unit abatement
ECONOMICS
4.917 4.883 4.819 4.789 4.664 4.276 3.956 3.892 3.758 3.712 3.694 3.665 3.635 3.583 3.567 3.544 3.475 3.379 3.376 3.321 3.296 3.245 3.226 3.223 3.220 3.216 3.190
Lignite-fired 232 234 240
Oil-fired 102 110 124 140 202
3.181
plants 0.439 0.438 0.436
4.743 4.735 4.712
228 253’ 280 300 342 345 420 480 613 630 660 1080 1900 1920 1932
2.698 2.630 2.571 2.534 2.47 1 2.467 2.386 2.340 2.270 2.263 2.252 2.160 2.098 2.097 2.097
42.655 41.591 40.656 40.071 39.066 39.004 37.733 37.002 35.893 35.785 35.607 34.161 33.180 33.166 33.158
plants 3.539 3.428 3.269 3.126 2.785
55.957 54.206 51.686 49.424 44.041
costs relative to SO, removed. costs relative to electricity output.
October
1988
291
Alternative
policies.for
sulphur dio.uide control:
H. Welsch
SO,, expressed in f/kg. For the quantity of SO, removed the maximum possible amount was assumed, ie 95 % of the original SO, quantity, according to the removal efficiency of the FGD process considered. The third column of the table contains the abatement costs per unit of output, expressed in f/MWh. It should be noted that the number of entries is smaller than the number of plants contained in the inventory because double entries of equal figures, stemming from identical plant and fuel characteristics were omitted. From the table it can be seen that both the unit costs relative to SO, removed (UC’)and the unit costs relative to electricity output (UP) differ substantially, depending on the fuel type. Within each fuel category UC’as well as uce increases as the capacity decreases. For coal-fired plants, UC’ranges from iO.27/kg for a plant of 2304 MW to f0.43 in the case of a 112 MW plant. The three lignite-fired plants have approximately the same capacity, requiring costs that range from f0.436/kg (240 MW) to fO.439/kg (232 MW). For oil-fired plants, UP is substantially higher in all cases, the minimum being f2.097/kg for a 1932 MW plant and the maximum &3.539/kg for 102 MW. To put these observations in perspective, it may be noted that, with the type of cost function given in Equation (6), UC”is always inversely related to capacity, operating hours and the SO, emission factor. (This can be shown by writing the amount of sulphur removed in terms of MW, h, sS and the removal rate, see Weizsacker and Welsch [21].) However, in our data, h, and sS are only distinguished by fuel type, not on a plant-by-plant basis. Moreover, a uniform, high value is assigned to h for both coal and lignite, whereas for oil h is lower by a factor of 10. On the other hand, E’for coal and lignite is only slightly different, whereas the corresponding value for oil is considerably higher, due to the higher sulphur content. These circumstances provide an explanation of the differences in ucs between fuel types, showing up in Table 1: The observation that UC”of the smallest coal-fired station is far smaller than UC’of the largest oil-fired plant reflects the fact that the difference in operating hours between solid fuels and oil is much larger than that in the corresponding emission factors and that the combined effect of these two differences outweighs economies related to capacity. The difference in UC”between coal and lignite fired plants, being caused only by a small difference in sS, is less pronounced. Regarding unit costs relative to electricity output, the picture is similar but not identical to the one just described. In particular, it is not universally true that coal-fired plants have lower UP than lignite-fired ones. However, for a given size (240 MW) UP is definitely lower in the case of coal firing. Regarding oil-fired
292
plants, it may be noted that FGD costs per MWh are in the same order of magnitude as average revenue of electricity delivered to industry. In order to assess the reliability of the figures obtained it might be desirable to check them against estimates from other sources. However, the information on FGD costs in the UK is rather scarce because, up to the middle of 1987, only one FGD unit had been installed in the UK electricity sector. The only useful source known to the author is Highton and Webb [7]. There, an estimate of 0.330 pounds per kg of SO, abated is reported for a power station to which FGD is retrofitted. According to Highton and Webb [S] this figure applies to a 2000 MW coal-fired plant with characteristics comparable to our assumptions. However, the estimate is based on an FGD process with only 80 % removal efficiency, in contrast to the 95 % assumed in the above calculations. Hence the amount of sulphur removed is smaller and economies of scale implied by set-up costs are less effective. Given this difference in the underlying assumptions and the technical progress that may have taken place since their estimate was made (eg the use of regenerative heat exchangers), their estimate seems to support our results. Of course, in view of the main purpose of this paper, the determination of SO, abatement costs on a plant-by-plant basis is only an intermediate step. As outlined earlier, the estimates of plant specific costs per unit of SO, removed serve as a basis for the construction of an aggregate cost function of SO, abatement in the British electricity sector. Table 2 shows, in the first two columns, the results of cumulating the quantities of SO, removed as well as the corresponding costs, according to the merit order of unit removal costs. The set of pairs of these two cumulated variables provides the aggregate SO, abatement cost function. Note the sharp cost increase in the range of about 2.59 x lo6 t of SO, to be removed. This marks the move in the merit order from solid-fired plants to oil-fired ones. In the marginal costs, shown in the third column, this jump is even more striking. The aggregate cost function provides a basis for the cost comparison of the two alternative regulatory strategies for SO, control to be carried out next.
Policy simulations The two types of regulation considered are uniform limits on the SO, concentration in flue gas (concentration standard) and transferable emission permits. The first is the regulatory tool most commonly employed in controlling SO, emissions (see OECD [ 131 for a compilation of concentration standards valid in various countries). The UK, however, has up till
ENERGY ECONOMICS
October 1988
Table 2. Minimum
aggregate
abatement
costs as a function
of SO2 removed. Marginal
Marginal SO, removed (1 x 103) 1SO.239 276.742 403.768 534.183 664.599 782.495 906.912 1031.785 1121.838 1209.477 1290.335 1365.454 1437.183 1499.392 1562.643 1623.287 1684.713 1740.792 1787.350 1830.387 1873.424 19 16.462 1959.499 2002.536 2045.573 2082.090 2118.867 2118.080 2177.293 2206.507 2232.068
Absolute costs (f x 106) 40.552 74.892 109.368 144.731 180.095 212.182 245.975 279.888 304.686 328.853 351.245 372.135 392.137 409.648 427.43 1 444.532 461.838 477.744 491.158 503.650 516.143 528.635 541.128 553.620 566.112 576.898 587.752 596.626 605.500 614.374 622.292
COSk3
%%)
0.271 0.271 0.271 0.271 0.272 0.272 0.272 0.275 0.276 0.277 0.278 0.279 0.281 0.28 1 0.282 0.282 0.284 0.288 0.290 0.290 0.290 0.290 0.290 0.290 0.295 0.295 0.304 0.304 0.304 0.310
SO2 removed (t x 103) 2256.586 2281.105 2304.97 1 2326.881 2347.8 13 2367.896 2387.459 2405.848 2421.498 2436.104 2450.711 2465.318 2476.272 2487.227 2495.639 2503.464 2511.159 2518.592 2526.026 2533.459 2540.763 2555.555 2569.977 2584.276 2599.457 2614.732 2629.754 2638.293 2643.511 2648.729 2653.948
Absolute costs (f x 106) 629.937 637.582 645.057 652.019 658.726 665.210 671.558 677.599 682.923 687.973 693.024 698.075 702.170 706.265 709.694 712.970 716.212 719.385 722.558 725.73 1 728.871 735.316 741.631 747.903 779.743 811.774 843.294 861.741 873.492 885.242 896.993
now pursued a policy of dispersing emissions by means of tall stacks rather than abating them, implying that a large part of SO2 emitted is deposited outside its boundaries. Taking for granted that this necessitates a certain overall reduction in UK emissions, the question arises how such overall reduction targets are best being implemented. The variability of FGD costs per unit of SO, removed, as discussed in the previous section, indicates a large potential for saving costs by employing cost-effective regulatory strategies instead of uniform emission limits. In what follows, the extent to which advantage can be taken of this potential will be examined. The method used is to compute the costs ofachieving various overall reduction targets by means of either concentration standards or transferable emission permits. The simulations presented are partly based on reduction plans currently aired. One is the agreement on a 30% reduction relative to 1980 emissions mentioned in the introduction. Another is the 60 % reduction, also based on 1980 emissions, according to the proposed EC Directive on the limitation of emissions from large
ENERGY ECONOMICS
October 1988
CO.@ wk)
0.312 0.312 0.313 0.318 0.320 0.323 0.324 0.329 0.340 0.346 0.346 0.346 0.374 0.374 0.408 0.419 0.421 0.427 0.427 0.427 0.430 0.436 0.438 0.439 2.097 2.097 2.098 2.160 2.252 2.252 2.252
SO, removed (t x 103) 2659.166 2664.384 2669.602 2674.821 2679.802 2684.648 2689.495 2693.290 2696.611 2699.338 2702.042 2704.414 2706.786 2709.000 2711.000 2712.803 2714.400 2715.507 2716.614 2717.721 2718.828 2719.808 2720.678 2721.484
Absolute costs (f x 106) 908.743 920.494 932.244 943.995 955.267 966.268 977.269 986.150 994.074 1000.802 1007.482 1013.493 1019.503 1025.195 1030.457 1035.319 1039.768 1043.227 1046.687 1050.147 1053.606 1056.811 1059.792 1062.646
Margioal costs (wz) 2.252 2.252 2.252 2.252 2.263 2.270 2.270 2.340 2.386 2.467 2.470 2.534 2.534 2.571 2.630 2.697 2.785 3.125 3.125 3.125 3.125 3.268 3.428 3.539
combustion plants. For each of these reduction plans, that uniform concentration limit which is necessary to achieve the given reduction target is derived. Note that the EC Directive, in addition to the 60% reduction rate, contains explicit concentration limits, namely 400 mg/m3 to be applied as from 1 January 1985 and 250mg/m3 applying after 31 December 1995. In the case of the UK, these are stricter than necessary for a 60% emission reduction. Therefore, a second set of simulations is run, based on the reductions resulting from these concentration limits. Finally, in order to provide a sensitivity analysis, various notional standards, specifying reduction rates relative to projected annual emissions in the reference period, are considered. The results concerning costs were obtained by modelling the emitters’ reaction to regulation in the way described earlier. This means that in the case of concentration standards the fraction of flue gas actually to be subjected to desulphurization is endogenously determined as a function of pre-abatement concentration, removal efficiency and concentration standard,
293
Alternutive
policies,fbr
sulphur
dioxide
control:
H.
Welsch
according to Equation (8). Overall abatement costs are then obtained by adding the abatement costs of all power stations. In the case of transferable emission permits it is assumed that the relatively most efficient plants (in terms of unit SO, removal costs) desulphurize their complete flue gas flow and the others take no abatement measures at all, the two subsets of plants being separated by the condition that aggregate SO, abatement conforms to the given reduction target. Overall abatement costs are then obtained according to the aggregate cost function tabulated in Table 2. The scenarios considered and the simulation results are summarized in Table 3. The leftmost part of the table contains the specification of the scenarios in terms of concentration standards iG,emission reduction rates p, the amount of SO, removed, R, and the quantity of post-abatement emissions, S. Beginning with the sixth column, the costs of achieving the removal targets R by means of either uniform concentration standards (CS) or transferable emission permits (TEP) are tabulated. C denotes total costs, UC” unit costs relative to SO2 removed and UC’ unit costs relative to electricity output. TCS, reported in the last but one column, is total costs saved by TEP, and CR is the percentage ratio of total costs under TEP and under cs. The first two scenarios are defined by concentration standards of 400 mg/m3 and 250 mg/m3 according to the EC Directive. The implied quantities of SO, to be removed are R = 2.548 x 106t and R = 2.667 x lo?, respectively. Relating these to the non-abatement emissions of 2.865 x 106t derived earlier gives abatement rates of p = 88.9 % and p = 93.1%. The postabatement emissions are S = 0.317 x 106t and S= 0.198 x 106t. Scenarios 3 and 4 refer to 30 % and 60 %
reductions relative to 1980 emissions. Since the latter amounts to 3.020 x lo’?, the implied reduction rates relative to the non-abatement emissions in the reference period are somewhat smaller, namely p = 26.2 and p = 57.8. The concentration standards that would be necessary to achieve these reductions are E = 2667 mg/m3 and 7t = 1524 mg/m3, respectively. Scenarios 5-11 are based on notional percentage reduction targets ranging from 30% to 90%. Turning to the results, the prime observation to be made is that there are in fact large cost savings to be achieved by TEP. Referring to reduction rates from 3&80%, total costs saved increase from 2208.795 to f279.890 million, while beyond 80% TCS decreases (see scenarios 1, 2 and 11). For p = 93.1% they are f 115.729 million, and it is evident that for the maximum overall reduction rate of 95% (as determined by the FGD process specification) costs under TEP are equal to those under CS, because in this case the allocation of abatement activities is identical under both regimes. Regarding the relative cost advantage, it can be seen that the cost ratio increases as reduction targets become more demanding, approaching, of course, 100 % as the reduction ratio approaches its maximum. Note that this increase is relatively flat over much of the range and becomes steeper only for reduction rates higher than about 90%. However, considering the proposed EC Directive these high reduction rates are politically very relevant. From the table it can be seen that overall reductions equivalent to the uniform concentration standard of 400 mg/m3 required in the Directive for 1985-95 can be achieved under TEP at costs that are 27.1% lower than under CS (scenario 1). In terms of total cost savings, the figure of f273.63 1
Table 3. Overall SO2 abatement costs resulting from policy simulations. Specification of Scenarios No I; P (m&m? W) 1 400 88.9 2 250 93.1 3 2661 26.2’ 4 1524 57.Sb 5 2530 30.0 6 2165 40.0 7 1807 50.0 8 1446 60.0 9 1084 70.0 10 123 80.0 11 361 90.0 Note: For notations ‘Corresponds Vorresponds
294
Costs under CS
2667 751 1657 860 1146 1433 1719 2005 2292 2578
198 2114 1208 2005 1719 1433 1146 860 513 287
1047.974 419.090 716.487 454.770 548.805 642.841 736.876 830.911 924.947 1018.983
0.393 0.558 0.432 0.529 0.479 0.449 0.429 0.414 0.404 0.395
Costs under TEP UC’ (f/MWb) 4.293 4.463 1.785 3.052 1.937 2.337 2.738 3.138 3.529 3.939 4.343
Coat comparison
F3;.;p,^’
;;!I
932.244 212.182 461.838 245.975 328.853 392.137 477.744 566.112 645.057 741.903
0.349 0.271 0.214 0.271 0.272 0.273 0.274 0.277 0.280 0.289
$kWb) 3.132 3.970 0.904 1.967 1.084 1.401 1.670 2.035 2.411 2.141 3.185
:“x 106) 273.63 1 115.729 206.908 254.650 208.795 219.952 250.703 259.132 264.799 279.890 271.080
:‘, 72.9 89.0 50.6 64.5 54.1 59.9 61.0 64.8 68.1 69.7 73.4
see text.
to a 30% reduction relative to 1980 emissions. to a 60% reduction relative to 1980 emissions.
ENERGY
ECONOMICS
October 1988
Alternative
million is close to the maximum of TCS attainable under all scenarios. Reductions resulting from a uniform standard of 250mg/m3, as envisaged for the time after 1995, would still allow for cost savings of 11 %, totalling 0 15.729 million (scenario 2). The differences in total costs between the two types of regulation are of course reflected in unit costs. However, it should be noted that the figures for UC under TEP are not exactly the product of their counterparts under CS and the ratio of total costs (CR). This is due to the fact that UC” under TEP is based on total SO2 reductions which are generally higher than those under CS. The reason for this is that, for simplicity, it was assumed that the marginal plant taking abatement measures under TEP performs desulphurization at full degree. It is not generally clear whether the figures given for UC under TEP overestimate or underestimate those valid for exactly matching the overall reduction targets. However, with respect to absolute figures, the above simplification implies an overestimation of total costs to be spent for reaching a given target under TEP. On the other hand, costs under CS are probably underestimated because a precise calibration of the fraction of flue gas to be subjected to desulphurization has been assumed. In practice, it is very likely that the desulphurization unit will be more or less oversized for reasons of uncertainty. Thus, the estimates of total cost savings seem to be on the safe side. It remains to comment briefly on desulphurization costs per unit of electricity produced. For high reduction rates (7&90%) the order of magnitude of UC” under CS is roughly one tenth of the price of electricity. Under TEP it could be about 30% lower, ie 7% of the electricity price. For 30% abatement the price increase induced by FGD is about 5% under CS, while it could be as low as 2.5%.
Conclusions This paper has been concerned with the costs of achieving overall reduction targets for sulphur dioxide emissions of the British electricity sector under two alternative regulatory strategies: uniform limits to SOZ concentrations in flue gas and transferable emission permits. The motivation for this investigation lies in the fact that the UK, being the largest SO, producer in Western Europe, would be particularly affected by political initiatives currently taken to reduce SO* emissions all over Europe. One such initiative is the proposal for an EC Directive on the limitation of emissions from large combustion plants. It contains the objective of reducing SO2 emissions by 60% relative to the 1980 figures and specifies a concentration standard of 400mg/m3 applying from 1985-95. For
ENERGY ECONOMICS
October 1988
policies jar sulphur dioside
control:
H. WeDch
the UK, this concentration limit would imply an emission reduction rate of 88.9% relative to annual emissions in the early 1990s. The question examined in this paper is, to what extent such reduction rates could be achieved in a more cost-effective way by employing the incentive mechanism implied by the polluter-pays principle. Note, in this context, that in the UK local effects of SO, emissions are alleviated by means of tall stacks. In order to control distant effects, on the other hand, it is sufficient to concentrate on overall emissions. The analysis is directed to the power plant sector mainly because it has a major share in the SO, emissions in the UK and because the concentation limits mentioned above explicitly refer to it. Regarding the results, the first statement to be made refers to SO, abatement technology. The estimates obtained indicate that, in recent years, improvements in flue gas desulphurization technology (which is the only one relevant to achieve reduction rates in the range considered) seem to have induced a certain decrease in the unit cost of SO, removal to be borne by individual plants. Second, relating to regulatory strategies, the properties of the FGD cost function imply that under transferable emission permits a given plant either performs desulphurization on maximum scale or takes no abatement measures at all. The cost savings obtainable from such a reallocation of abatement activities relative to a regime of concentration limits are in the range of 5&30% for emission reduction rates of 30-80%. Thus, percentage cost savings are inversely related to reductions in emissions. In absolute terms, however, cost savings increase from E209 x lo6 to ;E280 x lo6 as reduction rates increase from 3(X80%. Beyond 80%, not only percentage but also absolute costs saved decrease. Nevertheless, they are considerable. For example, the 88.9% reduction implied by the concentration limit of 400mg/m3 contained in the EC Directive could be achieved, under transferable emission permits, at costs that are 27.1% or almost E274 x lo6 lower. Of course, these estimates are only rough projections of the most likely outcomes of the policies considered. This is mainly due to the fact that differences in the efficiency, operating hours and characteristics of fuels burnt by individual plants could only incompletely be accounted for, due to lack of data. However, since the cost saving potential of transferable emission permits relative to uniform emission limits is based precisely on such differences, the estimation bias is probably on the safe side. Thus, it may be concluded that costs to be saved by the British electricity sector are considerable. Whether this justifies a step away from traditional regulatory patterns can, of course, not be decided on these grounds alone.
295
Alternative
policies for sulphur dioxide control: H. Welsch
The results concerning aggregate abatement costs are not only relevant for the decision on regulatory strategies to implement overall reduction targets but also for setting targets in the context of a coordinated SO, abatement policy on a European scale. Note in this context that the aggregate cost function derived is characterized by increasing marginal costs over the whole range, showing an upward jump at that removal quantity for which the desulphurization of oil-fired plants becomes necessary. The latter would be the case for the overall abatement implied by a concentration limit of 250mg/m3, as envisaged in the EC Directive for the time after 1995, but not for the reduction associated with a limit of 400mg/m3, being relevant for the early 1990s. As a consequence, the corresponding increase in the quantity removed of less than five percentage points would require an increase in minimum costs of more than 20%. This seems worth taking into account when setting emission targets for countries participating in an internationally coordinated SO, control policy. Alternatively, it could provide an incentive to smooth the cost function by changing the fuel mix. In either case, transferable emission permits should not be excluded from consideration when thinking about how to reduce the effects of British SO, emissions in a cost-effective way.
6
7
8
9
10 11 12 13 14 15
16
17
18
References S. E. Atkinson and D. H. Lewis, ‘A cost-effectiveness analysis of alternative air quality control strategies’, Journal of Environmental Economics and Management, Vol 1, 1974, pp 237-250. G. T. B. Camsey ‘Strategie der CEGB zur ErhGhung der Lebensdauer von Kraftwerken’, VGB Kraftwerkstechnik, Vol 64, 1984, pp l-5. CEGB (Central Electricity Generating Board), Statistical Yearbook 1984185. CIAB (Coal Industry Advisory Board), Environmental Strategies for Coal Use, Paris, 1985. EC (European Community), ‘Proposal for a Council Directive on the limitation of emissions of pollutants into the air from large combustion plants, Oficial Journal of the European Communities No C 49/l, 1984.
19
20 21
22
N. H. Highton and M. G. Webb, ‘Sulphur dioxide from electricity generation: policy options for pollution control’, Energy Policy, Vol 8, 1980, pp 61-76. N. H. Highton and M. G. Webb, ‘Pollution abatement costs in the electricity supply industry in England and Wales’, Journal of Industrial Economics, Vol 30, 1981. pp 49-65. N. H. Highton and M. G. Webb, ‘The effects on electricity prices in England and Wales of notional sulphur dioxide emission standards for power stations’, Journal qf Environmental Economics and Management, Vol 11, 1984, pp 7@83. L. Hordijk, ‘Towards a targetted emission reduction in Europe’, Atmospheric Environment, Vol 20, 1986, pp 2053-2058. IEA (International Energy Agency), Electricity in ZEA Countries, Paris, 1985. IEA, Coal Information, Paris, 1986. IEA, FGD-System Performance, Paris, 1986. OECD, Emission Standards ,f;)r Major Air Pollutants, Paris, 1984. Otto Rentz, Techno-okonomie betriehlicher Emissionsminderungsmafinahmen, Berlin, 1979. B. SchHrer and N. Haug ‘The cost of flue gas desulphurization and denitrification in the Federal Republic of Germany’, Economic Bulletin,for Europe, Vol 39, 1987. E. P. Seskin, R. J. Anderson and R. 0. Reid, ‘An empirical analysis of economic strategies for controlling air pollution’, Journal of Environmental Economics and Management, Vol 10, 1983, pp 112-124. T. F. Smith, ‘Factors affecting costs of power station emission controls with particular reference to the UK? in OECD, ed, Energy and Cleaner Air, Paris, 1986. UNIPEDE (Union Internationale des Producteurs et Distributeurs d’Energie Electrique), Programmes and Prospects for the Electricity Sector 1984-1990 and 1990-1995. 1985. United Nations, ‘EMEP: The co-operative programme for monitoring and evaluation of long-range transport of air pollutants in Europe’, Economic Bulletin for Europe, Vol 34, 1982, pp 2940. United Nations, Air Pollution Across Boundaries, New York, 1985. C. C. von Weizsgcker and H. Welsch, ‘Alternative Umweltstrategen fiir die europlische Elektrizitltswirtschaft: Zur Kontrolle grenziiber schreitender externer Effekte von SO,-Emissionen, in H. Siebert, ed, Um~veltschut-_,fir I,@ und Wasser, Berlin, Heidelberg, New York, 1988, pp 97-130. H. Welsch, Regulatory Strategies to Control Transhoundarx Sulphur Dioxide Pollution, forthcoming.
Appendix Table 4. Basic data for estimation
of SO, emissions
r~’sulphur content [X/100] cu (net) calorific value (MJ/‘kg] d sulphur retention factor [%/lOO] urn multiplier [kg/kg] e power plant efficiency [%/lOO] h operating hours [h/a] E*SO, emission factor [kg/M Wh] E”flue gas emission factor [m”/MWh]
296
Coal 0.016
23.4 0.05 1.9 0.369 5700 12.042 3363.391
Lignite 0.01
8.4 0.30 1.4 0.369 5700 11.382 4083.851
ENERGY ECONOMICS
Oil 0.029
42.25 0.00 2.0 0.287 500 16.645 3363.676
October 1988
Ahrnutiw
policirs~for
sulphur
dio.rid(> control:
H. l+ih-l~
Table 5. Basic data for estimation of FGD costs (Cf Equation (4)). annuity factor retrofit factor regression
coefficient
hourly labour costs annual labour requirements price of electricity unit input of electricity price of water unit input of water price of limestone unit input of limestone exchange rate Prices and exchange
ENERGY
0.17 1.33 f5340655.22 f8.80858 f2349.6775 f5 4000 hours f36/MWh 0.31 10-s MWh/ms flue gas 1 DM/ms 0.369 lo-“ m3/m3 flue gas fO.O152/kg 1.56 kg/kg SO, f0.278/DM
rates refer to 1985.
ECONOMICS
October
1988
297