Limits to costing the chronic health effects of sulphur components and other air pollutants

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LIMITS TO COSTING THE CHRONIC HEALTH EFFECTS OF SULPHUR COMPONENTS AND OTHER AIR POLLUTANTS P. A. WESTand C. DU V. FLOREY Department

of Community

Medicine,

(First receiced 7

St Thomas’

Hospital

Medical

School,

London

SE1 7EH. U.K.

December 1982 and receiued for publication 15 July 1983)

Abstract-Attempts to cost the effects on health of a variety of outdoor air pollutants have depended on epidemiolowal data. The difficultiesand limitations of the data and the costing procedures are discussed and Ilkrated. -

EPIDEMIOLOCv Epidemiological studies have provided the principal evidence of an effect of ambient air pollution on health because they are concerned with assessments of the population exposed to measured levels of pollutants under everyday conditions. Laboratory studies on humans exposed to known levels of pollutants have also given some information on acute health effects, but few of these studies have been done using the low concentrations of pollutants commonly found in ambient air. Many of these studies have been repeatedly reviewed over the last decade (Ferris, 1978; Holland et al., 1979; Rail, 1974; U.S. Department of Health, Education and Welfare, 1970) usually with an eye to establishing the levels a: which pollutants appear to have a detectable effect on some aspect of health. It is not our intention to provide yet another review but rather to consider the weaknesses ofepidemiological data for the purpose of economic analysis. The health outcome measures used in the epidemiological studies of long term effects of air pollution may be broadly categorized as mortality, morbidity and reduction in function within the normal range. Morbidity measures include the frequency of symptoms, especially those of the respiratory system, sickness absence and hospitalization. The costing of symptoms is not feasible in the absence of detailed information about the economic consequences of those symptoms. For children, the short-term consequences will fall mainly on the responsible adults and may be difficult to measure because of their diversity. The long-term personal economic consequences are not known. For adults the consequences will depend on sickness absence and hospitalization. Both of these measures are infrequently used in epidemiological studies because reliable data are difficult to obtain and routinely collected data are prone to substantial diagnostic error. The use of morbidity studies in the costing of health effects of air pollution is thus unpromising at the least. The costing of deviations within the normal range, such as minor changes in lung function, is yet 607

more difficult, since no impairment of activity would be expected. The best candidate for analysis would seem to be mortality, as there is a clear-cut outcome which may be related to lost productivity and use of health care. Yet even this outcome is not easy to handle either in epidemiological or economic terms. To illustrate the epidemiological problems, we have taken the studies of Lave and Seskin (1977) who estimated the cost of mortality attributable to air pollution in the United States. The general problems of economic analysis in the field are then discussed in detail. Mortality

The study of the relation between mortality and air pollution was stimulated by the massive number of deaths attributed to an acute episode of smog in London in December 1952. There followed studies of the acute effects of daily changes of pollution in London and in other cities from which the WHO Expert Group (1972) and the WHO Task Group (1979) concluded that increases in deaths were evident when 24-h average concentrations of smoke exceeded 500 pg m - 3 together with SO2 above the same value. More recent analyses of past data from London have suggested that variations in daily mortality may have been related to variations in daily. levels of smoke at rather lower levels, but problems with confounding factors and uncertainty as to which mathematical model is most suitable for describing the data prevent definitive statements about the lowest levels at which smoke and SO, are associated with acute increases of mortality (Mazumdar et al., 1982). The most often quoted analysis of the long-term effects of air pollution on mortality is that by Lave and Seskin (1977) using data from the United States. The relation was examined between mortality from all causes and certain specific causes in selected standard metropolitan statistical areas and air pollutants measured in these areas. The results of this study were used by the Organisation for Economic Co-operation and Development (OECD, 1981) to estimate the costs

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WEST

and C. DL‘ V FL.ORFY

of ill health due to sulphur oxides and, in particular, sulphates. Estimates were made of the number of deaths to be expected for each pgm--’ increase of sulphate in the air and from this was extrapolated the cost of morbidity due to sulphate. These analyses undoubtedly imply that air polluted with sulphates may be life threatening and are interesting for the hypotheses they generate. But there is no obvious causal mechanism nor is the temporal relation tested. If atmospheric sulphates really cause death, we do not know whether mortality is affected by contemporaneous levels of sulphate or by sulphate levels experienced over a period of years. From the age specific analyses the effect of sulphate seems greatest in the elderly which might occur with either temporal model. Epidemiologically these mortality analyses may be useful in defining potential new problems, but the ‘dose-response’ curves obtained from them portray neither dose nor response as defined for similarly named curves obtained from carefully designed and controlled laboratory experiments. The investigator has no direct control over which people are exposed to what specified doses, nor over the environment in which people live, and he can obtain only a rough estimate of dose from ambient air monitors using rather inaccurate methods of measurement. Thus, the precise form of the dose-response curve cannot be predicted under these conditions nor can it be certain, because of the lack of control, that some other unmeasured factor (or factors) related both to mortality and pollution levels is not the cause of an observed association. Thibodeau et al. (1980) reviewed Lave and Seskins’ work in detail. In a reanalysis of the data they showed that with the removal of a few observations with outlying values for mortality, the regression coefficients for sulphate and particulates changed markedly. that is. the coefficients were unstable. Other authors (Cracker et al., 1979; Lipfert, 1978) have reported different results using U.S. mortality data for somewhat smaller areas. Although the results have been disputed (Chappie and Lave, 1983) the failure of some to find a relation between mortality and sulphate levels shows how data drawn from the same sources may be analysed to give contrasting interpretations. Recent evidence from the U.K. (Chinn et al., 1981) suggests that current levels ofair pollution are unlikely to have detectable long-term effects on mortality. A wide range of causes of death in 1969-1973 was examined in relation to smoke and SO, measured in the County and London Boroughs. Comparison with similar analyses done for the period 19481954 and 1958-1964 indicated a decline over the 20-y period from strong to no association between pollution and mortality. Long-term mortality studies seem to offer either unstable estimates of strong positive association between mortality and pollution or, with modest changes in the specification of the model, little evidence of an association. Thus, with such uncertainty in the es-

timated effects of air pollution, one might question the validity of assessments of the associated costs even using so definite a measure as death.

COSTING HEALTH DAMAGE

In spite of the inappropriateness of statistically estimated dose-response relation, the relations are frequently used to generate estima.tes of the health damage due to pollution or the health gains that would follow pollution reductions. As well as the limitations of the dose-response curves on which the estimates are based. there are further deficiencies in the methods used to convert health effects into financial costs. It IS possible to convert an estimate of health effects, m terms of years of life lost and days of sickness absence from work. into a notional amount of money by the use of some set of monetary health values. But valuing health is highly controversial in itself as life and health may be of incalculable value to a single individual at risk. Viewed from a single individual’s position, the entire basis of the calculation may appear invalid. However, in practice, individuals as consumers, workers and tax-payers face many decisions in which the intended or subsidiary outcome is a change in health. In neither role do they appear to behave as if life and health were worth protecting at all costs so the objection that health has infinite value is misconceived. An obvious and widely used source of money values is the labour market. Workers are paid for the use of their time when in good health. Wage rates also influence payments made by social security systems to the sick. Since the majority of workers would presumably not be prepared to sacrifice their health completely for the going average wage, we may arguably be confident that a year’s wages or sick pay is a minimum value for a life-year in good health. Valuing health by earnings For all its appeal as a simple underestimate of the economic value of health damage, the earnings method suffers from a serious weakness. Although it may understate the value placed on their own health by individual workers (or non-workers), it does not follow from this that individuals will value all reductions in society’s health risks at a constant level. Consider medical technology as an example. It holds out a hope of life to almost every patient. However. all possibly advantageous techniques could only be provided to every sufferer if a very large proportion of national output were devoted to health services. Given the widespread public and political concern with health budgets approaching 10 per cent of national income, it seems implausible that people would value the health of unidentified members of society so highly as to bear the cost of all possible health services, no matter how small their contribution to health. It follows that, at the margin, the earnings basis for the valuation of health effects in money terms may

Limlrs 10 costmg the chronic health effects of sulphur components and other air pollutants

the value which society places on the changes in the general risk of mortality and morbidity from air pollution predicted by the dose-response relationship. Against this, it has been argued, for example by Broome (1978), that the low marginal valuation of risk lo unspecified individuals by a society is due to ignorance of the identity of the person who will suffer the projected health damage. Given such knowledge, however, the affected individual might be prepared to pay a high price to avoid the anticipated illness or death. By levying taxes to finance every possible riskreducing programme in the health field, governments would be protecting society from its ignorance. Yet, as long as the choices are made without certainty as to who precisely will suffer as a result, it is unlikely that society will accept the taxation required. Even if earnings are accepted as an appropriate basis for the valuation of health changes, the estimates of losses due to illness should not be seen necessarily as a loss of real production. The loss of production due to sickness is almost always purely notional, particularly in the present economic climate. Restoring a worker to health or saving his life will merely frustrate the employment of another worker. No change in total production is likely lo occur as a result. A casual observer might easily assume from the available estimates of health damage that a reduction in the hazard would generate the change in production implied by the earnings figures. This creates the illusion that an investment of society’s resources to reduce the damage to health will be self-financing. That is, it might be assumed that the cost of preventing the health damage will be offset by the extra production of the now healthier workers. This is unlikely to occur. In consequence, it cannot be assumed from estimates of the damage due to a health hazard and from the costs of the eradication of the hazard that eradication is necessarily a good investment. The effects of expenditure on reducing other health hazards must first be estimated. Only then will it be possible to rank air quality standards appropriately as a priority for society’s expenditure. For example, it may prove more cost-effective to equip every home with fire detection equipment or to enforce stringently legislation on the use of car seat belts than to improve air quality. Policy decisions must rest on an analysis of all programmes offering social benefits or at least all those with a direct health objective, because each such programme cannot be self-financing in terms of production gains. overstate

Efects on health sercices

Health on health However, economic very.least countries payment, provided

services are affected by any pollution effects which alter the demands for medical care. the decision to demand health care is an one. Patients incur costs, which are at the the time and trouble of attendance. In many health service use may also involve some depending on the type of insurance cover or purchased. Therefore, for a comprehen-

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sive analysis of the effects of pollution of health services, a range of variables, including prices, time costs and the local availability of health services, will be required, together with relevant individual characteristics. Once the correct dose-response relationship linking use of health services and pollution has been reliably identified, the computation of crude health service costs from average costs per hospital case, etc. is comparatively straightforward. However, in practice, any estimated savings from reduced pollution may again be notional rather than actual. Health facilities are typically used wherever they exist independent of the underlying level of community morbidity. Therefore, any resources freed following a reduction of air pollution would almost certainly be used to treat other cases. This is not in itself a bad thing since it gives care to a group previously denied it. But it does mean that no financial savings will accrue to fund the improvement in air quality unless health facilities are closed and resources shifted to other uses. Health costs in Europe-a

recent example

The OECD (1981) has recently produced a set of estimates for health costs that would be avoided by a reduction in air pollution. These use linear dose-response curves drawn from the work of Lave and Seskin. To take some account of the uncertainty of empirical estimates of dose-response, the OECD analysis adopted maximum and minimum dose-response coefficients respectively five times larger and five times smaller than the calculated effect of pollution on health. The estimates are calculated for two scenarios of air pollution reduction. However. we concentrate here on the impact of the larger change in air quality, a 37 per cent fall in emissions of SO,. Mortality effects, calculated in terms of changes in life-expectancy, are not monetarized in the OECD study, because of the sensitivity of attaching a price to life. For the 11 European countries examined, the maximum benefit from the proposed pollution reduction is in no case more than 1.5 y. Using the estimated mean relationship, the increase in life expectancy would be less than 0.3 y and for the minimum estimate. 0.06 y. The morbidity estimates of the OECD study are monetarized for sickness absence and hospital admission. Both are projected from the estimated dose-response relationship for mortality on the basis used by Lave and Seskin, that a given percentage reduction in total mortality might be expected to be accompanied by a similar reduction in total morbidity when air quality improves. This is necessarily an assumption in the absence of detailed separate estimates of the dose-response relationship for morbidity. However, it is not a particularly plausible assumption. Consider a situation where pollution causes 5% of mortality. A halving of this mortality

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brought about by improved air quality would be equal to

a fall in total mortality of 2.5 %. The effect on total morbidity could be bigger or smaller than this, depending on the share of morbidity due to pollution. Since some diseases cause a great deal of illness but few deaths, it cannot be assumed that pollution causes the same proportions of mortality and morbidity. It is difficult to specify exactly the diseases which may be caused or exacerbated by air pollution. However, data compiled by Black and Pole (1975) for the U.K. and Lindgren (1981) for Sweden provide convenient examples of the importance of morbidity due to causes wholly unrelated to air pollution. These data also show that broad categories of disease, (e.g. heart disease, cancers and respiratory disease) linked by Lave and Seskin to air pollution, account for a smaller proportion of morbidity costs than of mortality. (A more detailed analysis would subdivide these disease groups still further in an attempt to trace the components linked to pollution.) We can begin with the assumption that when air quality improves, mortality and morbidity from diseases affected by pollution change by the same proportion. For the U.K., heart disease, cancers and lung disease accounted for some 58.9 per cent of total years of life lost due to premature mortality in 1972. Thus a fall of 1.7 per cent in mortality from these diseases would amount to 1 per cent of total mortality losses. However, inspection of the hospital admission data shows that a 1.7 per cent fall in admissions for the three diseases would constitute only 0.2 per cent of total hospital bed days provided to patients of all kinds. A fall of 1.7 per cent in sickness absence from the same three diseases would constitute only 0.5 per cent of total sickness absence from work. (This figure would be further reduced, to 0.4 per cent if absence due to acute respiratory infections, coughs, colds and influenza, were removed. Clearly, some such spells of illness will be exacerbated by pollution but others will be unrelated.) The data for Sweden for the period 19661975 show a broadly similar picture, with diseases linked to air pollution contributing more heavily to total mortality losses than to total morbidity. For both countries, it is the importance of non-fatal conditions unrelated to air pollution that accounts for the lower contribution of pollution related diseases to total morbidity, e.g. mental illness, muscular and skeletal problems, and pregnancy. Overall, the data suggest that the assumed correspondence between projected changes in total mortality and morbidity when air quality improves cannot be sustained. The change in sickness absence is monetarized in the OECD study at the rate of compensation paid to sick workers. The total is then referred to as income saved from reduced morbidity. If, as has been argued above, production is not affected, the ‘saving’ constitutes a transfer payment. If improved health means reduced turnover of jobs, then savings may be offset by payments to the unemployed. Even if this does not

DU

V. FLOREY

occur, the saving will still not be real. Average incomes cannot be affected if production is constant. Forecasts of ‘savings’ to the health sector and from reduced sickness absence using the mean doseresponse coefficient refer to 1985. For the U.K., for example, they amount to between $2.5 and $6.6 per head, according to the level of pollution reduction projected. It follows from the five-fold selection of maxima and minima that maximum benefits reported were $12.5-30 per head and the minimum SO.5and $6 per head. Simple deflation of these estimates in line with the comments above, namely reduction of health service effects by 75 per cent and of sickness absence effects by 50 per cent brings the combined total cost down by some 65 per cent. Comparable deflation for other European countries’ estimates would typically reduce total costs by more than 40 per cent because of the higher relative cost of most European health services. To summarize, the cost effects of morbidity are overstated, given the method of calculation, and are still not large per head of population. They also imply, perhaps unwittingly, that benefits will be saved when in practice no changes in total production or health service costs are likely. However, potentially useful health services might be freed by a reduction in pollution-related disease. Nothing in the available estimates provides us with a reliable projection of this effect, however, largely because of the difficulties of establishing an unambiguous dose-response relationship. CONCLUSIONS It is not possible to place changes in mortality and morbidity on a common monetary footing by any simple calculation. Such calculations tend to mislead by encouraging the scrutiny of costs and benefits in a single area, such as pollution control. Pollution control costs, if incurred, cannot be viewed as an obviously profitable social investment since the incurring of costs here may frustrate expenditure on more useful programmes. In practice, monetarization of benefits is unrealistic. Rather, policy-makers should attempt to identify health benefits and the costs of achieving them for as wide a range of health-improving programmes as possible. The valuation of public health, compared to other competing uses of society’s resources, will then determine which health programmes will be adopted. This is inevitably a political question and is scarcely illuminated by attempts to generate health benefit estimates in aggregate cash terms. Judgements of the importance of death vs illness, and the health of the young vs the old will be required. None of these judgements are avoided by the use of simple money values from earnings data or other sources. These mask the choices with spurious technicality and may mislead by implying changes in real output which will not occur. Once a variety of projects have been evaluated, it will

Limits to costing the chronic health effects of sulphur components

be possible to judge the priority of air pollution in its correct context. This will require a systematic comparison of the alternative health benefits that can be achieved from the available budget.

control

REFERENCES Black D. A. K. and Pole 3. D. (1975) Priorities in biomedical research. Br. .I. prev. sot. Med. 19,222-227. Broome J. (1978) Trying to value a life. J. Pub. &on. 9, 91-100. Chappie M. and Lave L. (1984) The health effects of air

pollution. A reanalysis. J. urban &on. (in press). Chinn S., Florey C. div., Baldwin I. G. and G&go1 M. (1981f The relation of mortaiitv in England and Wales 1969-73to measurements of air polmtion.‘j. Epidem. Communiry Hlth 35, 114-179.

Cracker T. D., Schultz W., Ben-David S. and Kneese A. V. (1979) Methods development for assessing air pollution control benefits-Vol I. Experiments in the economics of air pollution epidemiology EPA-600/5-79-OOla. Environmental Protection Agency, Research Triangle Park, North Carolina. Ferris B. G. (1978) Health effects of exposure to low levels of regulated air pollutants. A critical review. J. Air Poseur. Control Ass. 28, 482-497. Holland W. W., Bennett A. E. B., Cameron I. R., Florey C. du V., Leeder S. R., Sehilling R. S. F., Swan A. V. and Wailer R. E. (1979) Health effects of particulate pollution: reap-

dnd other

air pob.mnts

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praising the evidence. Am. J. Epidem. 110, 525-659. Lave L. and Seskm E. P. (1977) Air Pollution and Human Healrh. Johns Hopkins University Press, Baltimore. Lindgren B. (1981) Cosr of Illness in Sweden, 19641975. Swedish Institute for Health Economics and Liber, Lund. Lipfert F. W. (1978)The association of human mortality with air pollution: statistical analysts by region, by age and by cause ofdeath. PhD thesis. The Goodwin Watson Institute for Research and Program Development. Mazumdar S., Schimmel H., Higgins f. T. T. (1982) Relation of daily mortality to air pollutron: an analysis of 14 London winters. 1958/59-1971!72. Arch5 ennr. Hith. 37, 213-220. Organisation for Economic Co-operation and Development. (1981) The costs and benefits of sulphur oxide control. Organisation for Economic Co-operation and Development, Paris. Rall I). P. (1974) A review of health effects of sulfur oxides. Enuir. Hlth Perspects 8, 97-121. Thibodeau L. A., Reed R. B.. Bishop Y. M. M. and Kammerman L. A. (1980) Air pollution and human health: a review and reanalysis. Emir. HIrh Perspecfs 34, 165-183. U.S. Department of Health, Education and Welfare (1970) Air Quality Criterta for Sulfur Oxides. National Air Pollution Control Admmistratton Publication No. AP-SO, US Government Prmting Office, Washington D.C. World Health Organi~tion (1972) Air quality criteria and guides for urban air pollutants. Technical Report No. 506. World Health Organization, Geneva. World Health Orgamzation (1979) Environmental Health Criteria 8 Sulfur oxtdes and suspended particulate matter. World Health Organization, Geneva.