This chapter discusses specific estimates of shadow prices (unit costs) for the damages from the impacts described in the previous chapter. Where uncertainty exists, a range of estimates is given. All estimates are in 1994 Canadian dollars unless specified otherwise. In applying the shadow prices estimated here, several cautions should be observed:

The character of an impact also determines its shadow price unit. Impacts from upstream activities (see Sections 3.1.1, 3.4.1 and 3.4.2) or waste disposal (Section 3.5), for example, can be represented per kilometre driven by a vehicle during its life. Correspondingly, the shadow prices for impacts from these activities can also be given per kilometre.
 
 

4.1.1 Fine Particulates

Available current studies by others show that inhalable particulates pose a major air pollution problem in urban areas from a human health perspective. Many studies have evaluated the damage to humans’ health from particle pollution. Examples of more recent studies include Hall et al (1992), and Friedlander and Lippmann (1994). The dose-response relationship in terms of the risk of illness or death at a given ambient PM₁₀ concentration probably does not vary much between cities. What does vary is the relationship between emissions, ambient pollutant concentrations and human exposure (Copes 1995).

Research undertaken for the ministry indicates that particulate matter is a significant local air pollutant, especially in terms of human health damage. We commissioned a study from SENES Consultants (1994) to provide an estimate of damage costs to the 1.8 million residents in the Lower Fraser Valley airshed. SENES based their assessment on parameter values for dose-response functions for human health impacts from Hall et al. (1992) and from Ostro (1993). Although these parameters are representative of Southern California conditions, SENES scaled the results of the analysis to reflect the levels of particulates in the Lower Fraser Valley airshed. A linear relationship between particulate concentrations and health impacts was adopted from recent literature. SENES allocated the damage costs to mobile sources based on 1990 emission inventory data, which is now under review. Finally, simplifying assumptions were made about the contribution of mobile source emissions to the formation of aerosols in the atmosphere.

Considerable uncertainty remains about the relative magnitude of source contributions to ambient levels of fine, inhalable particulates (PM_2.5) in the study area. Although not perfect, the screening-level estimates help to focus on those air pollution costs that matter most in transportation planning. According to the SENES (1994) study and subsequent revisions to estimates of vehicle contributions to ambient PM2.5 levels (Hrebenyk 1995), excess mortality caused by inhalable particulates emitted by mobile sources (including on-road vehicles, off-road vehicles, railroads, marine vessels and aviation) and generated as fugitive road dust in the study area is as listed in Table 4.1. The table assumes a C$3 million value of statistical life. The high and low estimates reflect uncertainty about the relative contribution of sources to ambient levels of PM2.5. Road vehicles contribute about 70% of the mortality shown in Table 4.1. Vedal (1995) estimated a lower bound on total PM2.5 mortality from mobile and stationary sources in British Columbia to be 82 people per year at present.

The statistics in Table 4.1 should be compared with about 150 fatalities, including pedestrians and bicyclists, who die annually in traffic accidents in the Lower Fraser Valley. At an average value of human life of C$3 million, these fatal accident costs are C$450 million. By comparison, mortality caused by fine particles from traffic, including road dust, is equivalent to about C$300 million to C$400 million per year.

In addition to the mortality costs, estimated identifiable morbidity costs (identified as the costs of minor restricted activity days) from fine particles are C$36 million (SENES 1994). Since it is plausible that other, yet-unidentified, types of morbidity are also attributable to inhalable particulates, this figure should be considered a lower bound on fine particle morbidity costs.

Extrapolating the Gothenburg estimate to the population of the Lower Fraser Valley would yield a total damage of about C$400 million from all sources. The SENES estimate was C$400 million to C$500 million for comparable types of damage from mobile sources alone, which contribute roughly half of the fine particulates. The discrepancy is probably due to better emission standards and enforcement in Sweden.

In a newer study, SENES (1995) estimated the PM2.5 mortality rates per kilometre travelled in various speed bands for different vehicle types and technologies, including future low-emission vehicles. These rates can be multiplied by the value of statistical life to arrive at fatality costs per kilometre travelled. A sample of PM2.5 mortality rates and fatality costs per kilometre for the speed of 40 km/h is presented in Table 4.2. Extrapolating from the SENES (1994) study, morbidity costs would probably add at least a further 10% to these values.

Senes (1995) gives mortality rates for the full spectrum of traffic speeds. The rates are flat for all speeds for diesel engines. For gasoline engines, the curve is U-shaped with respect to the speed variable. The range of estimates in Table 4.2 represents the uncertainty about the relative magnitude of source contributions to ambient levels of fine particulate matter in the Lower Fraser Valley. The high damage power of heavy-duty vehicles relative to light vehicles, and diesel vehicles relative to gasoline vehicles is obvious. Future vehicle technology will not drastically improve the fatality statistics, as the improvements in emission rates will be offset by increased vehicle travel.

Following the precautionary principle, the higher estimates in Tables 4.1 and 4.2, and in Senes (1995) should be used.

As discussed in Chapter 3, scientists agree that in the absence of policy intervention, the greenhouse effect will cause significant global warming by the end of the 21st century and will continue (IPCC 1995). The projected rate of warming and consequent climate change from existing conditions at the local and regional levels are uncertain. Estimates of damage depend on what warming scenario is assumed, which region is analyzed, and what particular impacts and costs are included. Any scenario builds on imperfect understanding of the interactions between biota, air, sea, ice and land with solar radiation, which together determine nature’s mechanisms for processing greenhouse gas emissions. There are complex feedback mechanisms, which are difficult to predict in the short time dictated by the needs of international policy decisions. Not enough is known about ocean dynamics, cloud formation, atmospheric chemistry or carbon uptake by plants to know definitely what exactly will happen and when.

Even if the "dose" of warming could be precisely estimated, the exact "response" of the natural systems affected cannot be calculated given current knowledge, either in terms of local climatic elements or in terms of the biosphere’s response. Extreme weather events will be more frequent and more intense than known from historical records. We can only guess whether and how pestilence and disease, for example, will propagate in the new climatic situations. Knowing which species will be harmed is not enough for estimating the costs of damage, either, because we are not able to evaluate reduced biodiversity and species extinction. Bein (1995) summarizes current understanding of possible damage resulting from global climate change.

Previous Estimates

One of the most comprehensive estimates of the costs of damage from global warming was made by Cline (1992). He calculated aggregate annual costs by category of damage in the United States for two scenarios:

Such large, non-marginal changes cannot be accurately priced. Nevertheless, for transportation project and policy appraisal, using a high shadow price for warming damage is appropriate, in order to ensure that decisions consider the worst-case scenario for global climate change.

Cline’s estimates are based on a relatively low value of human life (US$0.6 million). Recalculating the estimates with the value of human life of C$3 million, increases Cline’s monetized damage by 40% (from 1.1% to 1.5% of GDP, and from 6.1% to 8.5% of GDP for Scenarios 1 and 2, respectively). For reasons of international equity, human life lost as a result of global warming should be costed using the developed countries’ values, regardless of where in the world the fatalities might take place. IPCC Working Group III (1994) pondered the question of value of life without a conclusion.

To refine the calculation method by considering long-term population projections and associated CO2 emissions and growth in world GDP, the ministry commissioned a paper from Cline (1995a). Cline (1995a) examined a wide range of scenarios spanning a time horizon to the year 2300, and used the following parameters to estimate the shadow price:

For different combinations of the input variable values, Cline (1995a) calculated the cumulative tonnage of CO2 emissions, as well as the present value of greenhouse damage over all horizons stretching from the present into the year 2300, discounted at 1.5%. Shadow price of one tonne of CO2 is defined as the maximum ratio of the present value of damage to the cumulative tonnage.

Cline’s calculations concentrated on the "moderate-central" set of input assumptions and suggested that an appropriate shadow price in present value terms is on the order of US$20 per tonne of CO2. Cline’s version of weighting favours the moderate damage case, which is produced by a set of central values of the ranges he considered. The high-damage outcome (shadow price US$160 per tonne of CO2) was averaged out with a more moderate-damage case (US$70 per tonne of CO2) to represent a "high" damage case. A catastrophic-damage case (4.5°C warming at CO2-doubling, damage function exponent equal 3.0, and GDP loss of 2.5% and 4% in developed and developing countries, respectively) was not included in the weighting, although its shadow price is over US$300 per tonne of CO2.

These damage cost estimates are significantly higher than the widely used unit cost of carbon sequestering through afforestation (control cost). Although it has merit if used in combination with other measures, afforestation is not a panacea for long-term control because there is not enough land on earth to plant the number of trees needed to remove the volume of CO2 emitted by current human activities. Trees need time to grow and in the meantime the greenhouse effect will be active. Once a forest matures it is no longer taking up CO2, and releases carbon through forest fires and decomposition. Also, greenhouse gases other than CO2 cannot be removed from the atmosphere by trees, yet their future contribution may equal, or be greater than, that of CO2.

Alternative controls of greenhouse gases that have been proposed are too risky and too expensive to be considered for calculating shadow prices on the basis of control costs. Deflecting sunlight by seeding the stratosphere with aerosols or fine dust might do more damage than global warming. The estimated cost of scrubbing CO₂ from the emissions generated by fossil fuel combustion is astronomically high, and is likely higher compared to the damage costs of global warming (Schneider 1989).

The above shadow price estimates can be compared with the Swedish carbon tax, which is set at about C$60 per tonne CO2 (SEK333 per tonne CO2) and is being proposed to increase to about C$150 per tonne CO2 (SEK857) to stabilize the demand for fossil fuels at the 1990 level by the year 2000 (Kågeson 1995). In economic theory, a shadow price should be set at the optimal abatement level; a tax equal to the shadow price can then be used to internalize the externality. In practice, of course, the choice of a tax level depends on considerations other than the purely economic, such as how to manage demand for fossil fuels, and the political acceptability of taxation. In addition, carbon taxes in particular are not necessarily based strictly on the environmental costs of CO2 alone, but on all the costs associated with fossil fuel use. For these reasons, the Swedish carbon tax does not necessarily reflect the possible long-term damage from global warming.

Precautionary Estimate

The model of global warming damages presented by Cline is based on the longer-term, moderate damage case, which is produced by a set of central values of the ranges of inputs he considered. It represents the most probable scenario, with some weighting for risk. If nations co-operate and implement the abatement of greenhouse gas emissions effectively, starting immediately, then Cline’s model of global warming damage is likely to prove reasonably accurate. If business-as-usual continues, however, there are reasons to be concerned about the possible climate change and resulting damage from steady growth of greenhouse-gas emissions (Bein 1996). To examine the worst-case scenario, Bein and Rintoul (1996a), therefore, considered the effects of the "no abatement" case, using pessimistic assumptions about damage levels.

Bein and Rintoul used the Cline (1992) and Cline (1995a)model, with minor adaptations, but ran it over a longer horizon (to the year 2500), and with a set of more precautionary inputs to reflect a worst-case scenario (Table 4.3). They used the following inputs:

While the "no abatement" policy scenario represents the worst case, there are reasons to believe it could come to pass. One of them is the "free-rider problem" whereby any individual country has no incentive to take abatement measures, because if it did, it would lose (in economic terms) while other countries benefited from the abatement without enduring the costs. Commercial sectors dependent on the consumption of fossil fuels, such as the energy and fuel production industries, vehicle manufacturers and energy-intensive industries, may not support abatement actions. It is by no means certain that the human race will act to significantly slow down population and consumption growth. The probability of the worst case is not high, but it is real; the damages would be enormous and largely irreversible. Together, the probability and the damages mean the risk of the worst case is significant.

Bein and Rintoul (1996a) assumed 4% and 8% of GDP for the developed and developing countries’ benchmark damage, respectively. The estimate of damage from global warming, expressed as a fraction of GDP, is a crucial independent variable for shadow price estimation. A number of published estimates of the damage corresponding to CO2 equivalent doubling "agree" on about 1% to 2% GDP for developed countries, and 2% to 4% for developing countries (Pearce et al. 1995). Bein and Rintoul point out that this agreement does not necessarily demonstrate the reliability of the estimate for benchmark damage. There are several reasons for the similarity of the estimates:

The shadow prices depend strongly on the SRTP. A SRTP increase from 1.5% by 0.5% reduces the shadow price two to four times, with the strongest effect showing at the high bease damage and high damage function exponent. A reduction of the SRTP to 1.0% increases the shadow price two to four times. A high damage-function exponent affects the shadow price very strongly, while the degree of warming due to CO2 doubling (climate sensitivity parameter) has a weaker effect by comparison. The Bein and Rintoul study selects a shadow price of $1000 per tonne of CO₂ equivalent, favouring the peak price value over the average price and adopting a high value of damage function exponent out of precaution. The chosen value corresponds to high-damage with high-climate-sensitivity scenario at SRTP 1.5%. At SRTP 1%, the chosen value corresponds to low to medium-damage with medium to high-climate-sensitivity. These regions of the shadow price determinants have been avoided as unlikely by previous estimates.

It is evident that the choice of the shadow price must depend ultimately, both on assessments of what input values are reasonable, and on value judgments concerning how assertive the policy maker should be in establishing decision incentives.

Little cost data exists on human health, material and ecosystem damages due to increased UV-B radiation. The damages are most pronounced in the southern hemisphere, but the UV damage taking place there is an indication of things to come in the rest of the world. The cost data from that region is not firm because the research started only recently, and the transient nature of the ozone-depletion process and resulting UV-B radiation makes the task of establishing dose-response and physical damage relationships from observed data difficult.

The most studied damage is skin cancer melanoma in humans. Approximately twenty-six thousand new cases of malignant melanoma are diagnosed annually in the United States and 7800 deaths occur from the disease. Changes in lifestyle (more leisure time, more suntanning, less clothing), which involve increased exposure to sunlight, have resulted in more than a 200% increase in melanoma over the past 35 years in the United States, with a 150% increase in mortality (National Cancer Institute 1987).

Previous Estimates

In most medium-risk populations melanoma rates have doubled over the last 15 to 20 years (Elwood 1989). A major study for the United States Environmental Protection Agency predicted that several million additional deaths from melanoma would result in the United States over the lives of individuals alive now and born by 2075, if the ozone layer continues to deplete at 1980s rates (Longstreth 1989). This is roughly equal to the projected traffic fatalities in the United States over the same period. Counting the value of human life alone, the damage in the United States from melanoma would be on the order of US$100 billion annually, or about 2% of GDP. These risks are probably underestimated, as the rate of ozone depletion has increased since the time of that study.

In Australia, cortical cataracts account for the majority of projected morbidity and mortality costs due to increased UV-B radiation. A conservative estimate of total annual costs between 1995 and 2030 (including medical and non-medical expenditures and economic losses plus the value of human life, pain and suffering), are expected to be on the order of A$1 billion to A$2 billion annually by 2030 (Bryant et al. 1992).

The estimated upper bound of human health and mortality costs until the middle of the 21st century amount to about 1% of the current Australian GDP. By comparison, the United States estimate of mortality costs from melanoma alone (which in Australia accounts for only 9% of the increase in mortality and morbidity costs due to UV-B) is 2% of GDP. The damages to agriculture, forestry, building materials and objects made of fibres should be added to the estimates. Possible damage to basic life-support functions of ecosystems, and changes in biodiversity remain intangible and are potentially catastrophic impacts.

The loss of the ozone layer closer to the equator (mainly developing countries) may not be as significant as over countries located closer to the polar regionss, where the loss of stratospheric ozone is most dramatic at present. On the other hand, the developing countries are more dependent on natural resources that might be affected by the increased ultraviolet radiation. The developing countries would also be more vulnerable to changes in the ozone concentrations because they do not have as many resources required for successful adaptations.

It is prudent to assume on precautionary grounds that the long-term (middle of the next century) damage in terms of current GDP of the developed world amounts to 2% of GDP for incremental mortality and morbidity due to UV-B. The other impacts of UV-B radiation on the economy probably amount to another 2% of GDP, for a total of 4% of the developed world’s GDP annually. Since the cost data is scarce for developing countries, it is also prudent to make similar similar assumptions for those countries at this time.

The conclusion is that the total annual cost of damage from stratospheric ozone depletion, if calculated with the integrated costing method, would be close, in terms of GDP, to the order of magnitude of estimates for the cost of damage from global warming. The time streams of the costs due to increased UV-B radiation might not be as long as those for continuing warming and climate change, because ozone-depleting substances (ODSs) are targeted for a global phase-out by 2010. However, the possibility of surprises and ecocatastrophes is comparable, and therefore the longer-term intangibles might be damaging to a similar degree.

Precautionary Estimate

A preliminary shadow pricing of ozone-depleting substances (Bein and Rintoul 1996b) uses a method similar to that employed in shadow pricing of greenhouse gases (Bein and Rintoul 1996a). The shadow price is calculated as the ratio of present value of future damages to the cumulative emission over the analysis horizon. Discounting future damage costs to present values is based on ethical consideration of future generations, and it is termed "social rate of time preference" (SRTP), with SRTP ranging from 0% to 5%. The higher end of the range overlaps the discount rate based on the opportunity cost of capital, which is commonly used in cost benefit analysis of public projects and policy (IPCC 1995a). The on-going philosophical dabate regarding the choice of approach to discount rate is thus irrelevant at this end of the range.

A range of emission scenarios were assumed. Under the Optimistic Scenario, ODS emissions peak in 1995 and then drop to nil by year 2000. This scenario is purely hypothetical. The strictest reductions in ODS emissions negotiated so far under the Montreal Protocol will produce a scenario worse than the Optimistic Scenario. It is apparent that the currently set emission reduction targets may not be achieved. At the other extreme (the Pessimistic Scenario), ODS emissions continue for 50 years at the 1995 level and drop to nil by 2050. The actual future will thus likely fall between our assumed limits.

The benchmark damage was assumed to be 1.0% of global GDP in 1995. For other values of damage, the results can be scaled by proportion. The atmospheric life cycle of ODSs is assumed to be 100 years, based on a weighted average of 10 species of ODS lifespans from 6 to 550 years (WMO 1991).

The analysis is run over a range of SRTP, while ODS emisson futures range from the Optimistic to the Pessimistic Scenario. Table 4.5 summarizes the shadow prices for the range of assumptions. The shadow price is less sensitive to the SRTP than in the case of global warming, probably because of the shorter horizon over which ODSs are assumed to inflict damage, compared with greenhouse gases. Within each SRTP, the results are only moderately sensitive to the scenario chosen.

For the proponents of discounting based on the opportunity cost of capital, the shadow price value is about US$400 (C$500) per kg of CFC-11 equivalent.

The attribution of the above CFC damage costs to a unit of transportation work in project appraisals requires estimation of the mass of ODSs per vehicle from both vehicle manufacture and vehicle air conditioners, including recharging. The total mass of ODSs in the vehicle fleet at any future time strongly depends on the effectiveness of new regulations governing the automotive industry and ODS disposal and recycling. Per-kilometre unit costs must assume the quantity of ODSs consumed by transportation activities, and then spread out the consumption over annual vehicle kilometres and vehicle life.

This is straightforward for the existing situation, that is, the base case in social cost benefit analysis. Changes to the quantities of ODSs used by transportation and changes to future vehicle utilization are more difficult to predict, but the unit costs would strongly depend on the assumptions. Possible changes in travel behaviour, with less distance driven as a result, would increase the ODS contribution per kilometre. The total emissions depend on distance driven only to the extent that driving vibrations contribute to leakage and to the need to replenish CFCs. Even if left in a garage and never driven, a vehicle is a contributor to the ozone-depletion problem if it contains CFCs and if ODSs were used to make the vehicle and its parts.

For the base-case, average, light vehicle in British Columbia, the estimated annual mass of CFC emissions, including 10t from junked cars, is 245 t, which would be responsible for almost C$200 million in economic damage, assuming C$800 per kg of CFC equivalent for damage equal to 1% of global GDP. For 4% of world GDP as a measure of damage, the B.C. vehicle responsibility would be $0.8 billion. Damage from ODSs used in vehicle manufacture plus intangible damage would be extra. Assuming 16 000 km per year utilization in British Columbia (Bein et al. 1996), the damage distributed among all of the 1.3 million vehicles is $0.039 per kilometre plus intangibles. Calculated only for the vehicles containing CFCs in air conditioners (37% of all cars), the unit cost is about $0.11 per kilometre. These figures do not include the damage due to ODSs used to blow foam for vehicles, or any other ODSs used in vehicle manufacture.
 
 

4.1.4 Ground-Level Ozone and Precursors

Based on SENES (1994) and Rafiq (1986), annual crop damage in the Lower Fraser Valley area attributable to ground-level ozone amounts to about C$10 million (total annual agricultural production value about C$140 million). Human morbidity due to ground-level ozone (sore throat, cough, headache, chest discomfort, eye irritation, restricted activity days) costs some C$5 million annually. Ozone crop damage and human morbidity impose locally significant costs in outlying communities that receive the bulk of ozone generated in the Lower Fraser Valley. Reduced visibility from ground-level ozone has psychological and aesthetic impacts, and possibly economic costs to the tourist industry. Material damage costs other than crop damage have not been estimated. The cost of ground-level ozone damage is one order of magnitude smaller than the mortality costs of fine particulates.

In Swedish cities, Leksell and Löfgren (1995) estimated SEK48, SEK300 and SEK720 per kilogram of either nitrogen oxides (NOx) or volatile organic compounds (VOCs) for mean of all emission conditions throughout a large city, central street in mean weather and central street in light breeze, respectively. These values are based on willingness of Swedes and Norwegians to pay for a drastic reduction in pollution levels, and they are equivalent to C$8, $50 and $120 per kilogram, respectively. In a smaller town, these values would be half or less. For the city of Gothenburg (population about 450 000), the annual total, including particulates, NOx and VOCs amounted to SEK700 million (C$115 million), and for the whole country SEK8 billion per year (C$1.3 billion). These figures consist mainly of health costs.

Leksell and Löfgren also attempted valuation of damage by NOx and VOCs to ecosystems. They recommended a shadow price of SEK40 per kilogram of NO_x (C$6.50 per kilogram), based on the marginal cost of reducing industrial NOx emissions to abate acid rain. Data was lacking on costs of other forms of damage by NOx, which is also a precursor of ground-level ozone. This value has been used in Sweden since 1992 as an environmental charge, and with inflation, this charge would be SEK45 per kilogram of NO_x (C$7.30 per kilogram) today. The authors also recommended using SEK20 per kilogram of VOCs (C$3.25 per kilogram), another precursor of ground level ozone.

4.2.1 Traffic Noise

Many reasonable methods exist for calculating noise levels as a function of traffic speed, vehicle mix and operating mode for different layouts of road and roadside. However, costing of noise impacts lags behind. Although numerous cost estimates exist for transportation noise, most of them are limited to control cost and hedonic property value methods, which only assess a minor portion of the total cost, and thus understate total costs several times.

Estimates of noise impacts on human enjoyment of rural, remote and wilderness areas are lacking. So are studies of the effects of traffic noise on wildlife. In urban and suburban areas, one indication of noise impact is the reduction in the value of residential property adjacent to road traffic. Several studies show an average reduction in residential property values of about 0.5% for each unit change in Leq. (Leq is the equivalent continuous sound level in dB(A) (decibels weighted for human hearing sensitivities) over a set period.) Various researchers have used hedonic cost study results to develop general property value depreciation indicators, as shown in Table 4.6. The OECD recommends a noise depreciation of 0.5% of property value per decibel increase if noise levels are above 50 dB(A) Leq (24 hours) (Modra 1993).

From analysis of studies of depreciation of residential property values due to noise in OECD countries, Lamure and Lambert (1993) observed that depreciation was negligible if not nil in the 1960s. In the 1970s, it was between 0.3% and 0.8% per decibel increase in noise in the majority of studies. In the 1980s, especially in the second half, the rate was about 1%. The authors also show the results from a German study clearly indicating willingness-to-pay for noise abatement rising with the level of noise and with the income level of the person questioned. The willingness-to-pay also increased by an average of 20% with the level of understanding by the respondents of the damages caused by noise. These results might mean that as the awareness of traffic noise damage rises, better valuation methods are developed and the estimated values increase, and may indicate that further increases can be expected in the shadow prices of traffic noise.

The costs of traffic noise are reflected in expenditures made to reduce their impacts. As with hedonic pricing methods, such control costs greatly understate actual noise costs, because residual impacts remain even after control measures have been undertaken. Washington State Department of Transportation (1987) uses a formula for calculating maximum investments in noise reduction. The resulting values are shown in Table 4.8.

Table 4.10 shows the expenditures on the protection of population from traffic noise in selected Scandinavian countries.

Otterström (1994) reported an estimated C$1300 (1991 FIM 5000) per affected person per year as the damage cost of noise. Although it is not clear by what method the estimate was obtained, anti-noise expenditures in Finland are similar to the damage cost (Table 4.10). The calculation assumes that 33% of exposed people experience damage from noise levels between 55 dB and 65 dB. At noise levels of 65 dB to 70 dB, 50% of exposed people are bothered. All exposed people are affected above 70 dB.

It is not clear from any of the references whether the social costs of forgone opportunity to build residential developments and recreational areas along busy arterials and interchanges is included in the total figure. If not, the total cost of noise would be higher than indicated.

On a relative basis, Canadian heavy trucks cause about twice as much noise damage per tonne-kilometre hauled as do freight trains (Transport Concepts 1994). Estimates of damage per average person affected by truck noise compared with train noise are not available.

Based on values currently used in Scandinavian countries, a shadow price of C$1000 to C$1500 per affected person per year is adopted.. This is likely a lower bound for the total damage cost of noise.
 

4.2.2 Traffic-Induced Vibrations

Some estimates of damage from vehicle vibration assume that ground vibrations caused by large trucks and buses are responsible for a large part of urban building structural damage costs. This finding disagrees with Watts (1990) who studied the literature and performed vibration tests on buildings, without finding any significant link between traffic and structural damage in buildings. The costs of vibration impacts on people are not known, except that roughly the same proportion of residents are seriously bothered by vibration as those seriously bothered by noise (Watts 1990).

4.3.1 Review of Land Use Impact Cost Studies

This section reviews a range of sources that consider the value of external environmental benefits for various land uses. When road projects and vehicle traffic alter land uses, the resulting change in environmental benefits can then be valued. Some of the sources provide estimated values for several types of benefits, while others concentrate on only one or a few categories of benefits. Both the methods used and the values estimated are discussed.

There are several general methods used to measure the environmental value of land. Some approaches focus on the contribution that an environment makes in the marketplace, such as the role of streams in commercial fishery production and the tendency of greenspace to increase real estate values. The value of non-market goods, such as recreational activities, should also be included in assessments of environmental values. Other valuation methods focus on the uniqueness and resilience of environmental resources (King 1994), or the energy embodied in and produced by an ecological system (Odum 1994). Some analyses incorporate a combination of these methods to estimate the total value of an area of land (Mä ler et al. 1994).

Valuation of direct annual economic benefits supplied by wetlands is discussed first because it is the basic input for the land-use change valuation method presented in Section 4.3.2.

Wetland Benefits

Studies of the external environmental benefits of wetlands are summarized in Table 4.11.The Michigan wetland annual values are each based on only one external benefit, which explains their relatively low magnitude. The Louisiana tidal marsh annual benefits (biological plus social) add up to a value comparable to that in the last line in the table. The Southeast United States values are the present value of selected benefits of wetlands. The present value reaches US$17 000 (1983) per acre (C$86 300 (1994) per hectare) with a 3% social rate of time preference used to discount the stream of future benefits. The benefits included commercial fisheries, trapping, recreation, storm protection and energy productivity of wetland ecosystems, but did not include non-use benefits, biodiversity or preservation values. The values are then best interpreted as minimum estimates.

This example demonstrates that the market value of land may be entirely unrelated even to the minimum value of the ecological services it provides. In the case Farber (1995) examined, market values were much less than the environmental values, but the opposite can as easily be true: if wetlands were filled in and developed as residential properties in an urban area such as Greater Vancouver, they could sell for perhaps $1 million per hectare. The annual profit from rent collectible on residential units built on the land could easily exceed the annual loss of direct, measurable ecological economic benefits from the wetland. An important warning follows from this: since the existing economic valuation of wetlands as ecosystems cannot save them from development if monetized values alone are used as the basis for decisions, the evaluation framework must consider non-monetized and intangible values of wetlands as well for a complete analysis. When all costs and benefits are included, whether monetized or not, then decision-making can rely on a balanced tool for sustainable development evaluation.

These estimates indicate that typical minimum environmental economic benefits of wetlands are C$30 000 per hectare per year, including values of general recreation and ecological and hydrological functions. In terms of annual benefits, a Louisiana-type wetland provides some C$20 000 per year in primary biological production alone. The annual values would be higher for wetlands that provide unique benefits, such as a habitat to endangered species, a large wastewater treatment capacity or existence value to a large urban population. The value increases significantly if loss of the wetlands implies an irreversible loss to future generations.

Wetland Restoration and Engineering Costs

Restoration costs of damaged wetlands and costs of replacing a lost wetland at another location are an application of the control-cost method of shadow pricing. Studies cited by Kirby (1993) give estimates of the cost of wetland replacement ranging from $21 000 to $178 000 per hectare. These estimates only reflect the costs of modifying existing land uses and do not include land acquisition expenses.

Mitsch and Gosselink (1993) cite a number of wetland engineering projects in the United States. The costs per hectare of habitat created for wastewater treatment ranged from US$74 000 (1989) to US$2 million (1989), but these were typically small total areas. Urban runoff and stormwater runoff projects cost US$200 000 (1989) and US$52 000 (1989) per hectare of wetland, respectively. Habitat and recreation types of wetlands cost US$41 000 (1988) to US$70 000 (1987) per hectare to construct. The lowest cost of wetlands constructed was for a dredge disposal site at US$25 000 (1982) per hectare. The median (calculated by Mitsch and Gosselink without adjusting prices for different years) was US$74 500. This would be about C$110 000 (1994) per hectare. These costs are probably lower limits since the degree of success is not specified for the projects.

Forest and Tree Benefits

A forest can provide timber production; a place for tourism and recreation; habitat for fish, wildlife, rare species and special natural features; a water reservoir; existence, bequest and option values; andaesthetic, heritage and scientific values.

Some of these values (commercial value of timber, for example) are relatively easy to calculate. Tourism benefits are more difficult to determine. Annual B.C. tourism revenue was estimated at $3.2 billion in 1994. How much this tourism depends on forests is uncertain. Provincial travel promotion efforts to emphasize forests and outdoor activities that depend on forest habitat indicate a significant contribution.

The B.C. Ministry of Environment, Wildlife Branch (Reid 1986, Reid 1990) and Environment Canada (1993) have sponsored studies of consumptive and non-consumptive wildlife recreation in the province, including estimates of expenditures, net economic values and employment associated with these activities. These studies indicate that over 95% of B.C. residents participated in one or more wildlife-related activities, on which they spent a total of $977 million in 1991. An increasing portion of these activities are non-consumptive (Environment Canada 1993). Much of this wildlife depends on forest habitat.

A number of studies have identified the monetary value provided by individual trees in specific circumstances (Brabec 1992). Australia has developed a draft tree "amenity value" formula that is used to calculate the value of a particular tree in a developed area (Standards Australia 1992). The formula incorporates visual impact, frequency of occurrence, historical significance, form and vigour, crown size, life expectancy and site suitability. Total tree values range from a low of $71, to a median value of $6200, to a maximum value of $480 000 (1994 Canadian dollars).

Energy Productivity

Some scientists propose that primary energy productivity can be used to indicate the environmental value of an area of land (Odum and Odum 1976, Costanza 1980, Odum 1994). Although values such as diversity and aesthetics are not counted, the quantity of energy production is the basis for many factors that make an area environmentally valuable. This approach has been used by Costanza et al. (1989) in conjunction with other costing methods to monetize wetlands, and is a subcomponent of other valuation techniques (Hannon 1991, Ulanowicz 1991). This technique supports the general ranking of land types in Table 3.12 and should be useful for relative ranking of land use type values.

Greenspace Preservation

A number of surveys have used contingent valuation (CV) and hedonic pricing techniques to identify the value to society of preserving greenspace. For example, a study found the mean willingness-to-pay to preserve Birkham Wood in North Yorkshire from highway development to be £18.94 (C$38) per person surveyed (Hanley and Spash 1993). Other CV studies mentioned in Navrud (1992) and Carson et al. (1995) also show support for preserving various types of greenspace.

Walsh et al. (1984) surveyed Colorado residents concerning preservation values of wilderness. Respondents put a value of $694 per hectare for increasing the state’s 1 million hectares of wilderness to 1.5 million hectares. Approximately half of this benefit was based on "non-use" benefits, including option, existence and bequest values. This indicates that cost estimates which rely only on "use" values (recreation, tourism, extraction of resources) can significantly understate the total value of conservation. It shows the importance of including non-use benefits in environmental valuation. A study for the Bonneville Power Administration established the values for environmental resources shown in Table 4.12. It is not clear which values are based on contingent valuation and which are not.

A portion of greenspace external environmental benefits is reflected in higher property values of adjacent real estate. For example, urban land adjacent to protected farmland in Oregon was worth $5285 per hectare more than similar land 300 m away (Nelson 1986). A study in Boulder, Colorado found that housing prices decline with distance from a greenbelt. Houses adjacent to a greenbelt are worth an average of 32% more than an otherwise comparable house 1 km away (Brabec and Kirby 1992). Philadelphia’s 524 ha Pennypack Park increases nearby real estate values by an estimated $11 million, averaging more than $29 000 per hectare of park (Brabec 1992).

This technique can be used to identify some benefits of greenspace preservation. However, these hedonic pricing techniques are only able to capture a portion of the land’s total aesthetic and recreational value, since greenspace beauty and recreational benefits are also enjoyed by non-residents. Other environmental values, such as wildlife habitat, existence value and water quality are not measured either. For these reasons, estimates of external benefits of greenspace based on hedonic pricing should be treated as a lower bound on total values, and the non-monetized benefits not captured by hedonic pricing should also be considered.
 

4.3.2 Simplified Valuation of Land Use Impacts

Ideally, evaluation and monetization should be performed for the external value of each specific site individually. This is essential if the land has uniquely valuable features. Since this is often impractical, the second-best alternative is to develop generic value estimates of various classes of land use. Generic values are also needed as inputs to high-level planning.

An ideal model of generalized environmental costs for changes of standardized land use categories due to roads, motor vehicle traffic and development would be based on the following three steps:

In reality, each of the three steps above is complex and data-intensive. This ideal model will probably never be fully reached, because:

The first step ranks land use types in terms of their environmental benefits, based on the relative values described in Table 3.12. The ranking is by ordinal value, on the basis of the values and considerations raised in the literature, where available. The ranking is:

The third step constructs a linear value function for land uses relating value and benefits, using the value of pavement as the minimum and the value of wetlands as the maximum. For each type of land use between wetlands and paved land, an order-of-magnitude estimate has been assigned, based on its relative ranking among types of lands. Equal spacing along the curve is assumed for lack of more precise data. No absolute values are known. Figure 4.1 shows the resulting linear value function.

Table 4.13 shows the resulting default values in the form of the benefits gained or lost due to a change in land use. These estimates, intended for strategic and high level planning applications, are not absolute, and do not include the value of unique resources located or dependent on a particular piece of land affected by a road project, which should be assessed on a site-specific basis.

For each hectare of land that will be converted from its current use in the left column to another use defined in the top row, the dollar amount in the intersection cell indicates the change in ecological economic value. We assume that wherever a land use is changed from one type to another, the same gain or loss of benefit occurs. The valuation of land use change applies to land affected directly by a project, as well as land affected indirectly, such as residential development made possible by the road project.

For each hectare that will be disturbed, but not changed in type, arbitrarily half the value for conversion should be charged. Half is a central estimate between 0% and 100%. This allows for a degradation of land use, rather than a change: for example, if farmland is made less productive by the impacts of a transportation project.

Example: If a road project requires paving 20 ha of farmland and 10 ha of second- growth forest, leading to residential development on 10 ha of second-growth forest, and causing noise and pollution impacts to 5 ha of wetland, 20 ha of second-growth forest and 30 ha of farmland, the total losses of environmental benefits are $810 000 per year as shown in Table 4.14. Non-use values, indirect economic values and intangible values are not counted.

In summary, this method is based on using the best available estimates for some land use types to "anchor" a sensible ordinal ranking of other types. Despite its limitations, we feel the method will provide order-of-magnitude estimates for high-level, overview planning that are an improvement over omitting the values of land use changes on the grounds that they have not been fully quantified. We believe that the use of these first-cut values will improve the quality of transportation decision making.

One-Time Land Restoration and Engineering Costs

Evaluation of projects involving decommissioning and relocation of existing facilities, or ecological engineering to recreate habitat lost through new road construction should consider the one-time costs involved. The one-time costs should be in addition to the annual operating costs that may be required to maintain the constructed habitat at the intended function and to monitor its performance. If the annual benefits of a habitat are only lost temporarily because of development, and the habitat is subsequently rehabilitated, the cost of rehabilitation should be included in transportation development project appraisals. Likewise, replacement of a habitat lost to development at another location should also be appraised.

It is important to note that attempts at restoration or creating new habitats, especially the highly productive types such as wetlands, are not fully successful, and in any case, total restoration is unlikely following severe damage (Farber 1995a). Ecosystems are the product of a long history of climatic, hydrologic, biological and other conditions, which humans cannot simulate through construction. Also, many attempts at restoration or engineering are unsuccessful because monitoring and necessary human intervention cease too soon after the construction is completed. Restored and engineered habitats, like gardens, require tending and a helping hand until the habitat heals and grows robust enough to withstand naturally occurring shocks (Kent 1995). Appraisals, therefore, should duly allow for permanent partial loss of ecological benefits, for probability of rehabilitation and replacement success, and for the costs of monitoring and corrections.

Habitat restoration and engineering costs depend on the type of habitat and are site-specific. Sometimes the restoration effort required is modest (Morley et al.1995). If site-specific cost information is not available in high-level studies, the planner may use C$100 000 per hectare as restoration cost for wetlands (lower bound), as discussed in the section, "Wetland Restoration and Engineering Costs", earlier in this chapter For other habitats, the restoration cost can be extrapolated from that value.

Testing Values

The appropriateness of the order-of-magnitude estimates for land use values given in Table 4.13 can be tested by asking: What is society willing to give up to protect environmentally valuable lands from development?

Although a few years ago wetlands were frequently drained and filled, most jurisdictions now restrict or prohibit loss of wetlands. The United State’s national policy of "no net loss" of wetlands implies that wetlands have a high marginal value. Investment over the next 20 years of C$6500 to C$20 000 per hectare of "rescued" Louisiana wetland will be equivalent to about 10% of fisheries benefits from these wetlands alone. In other words, C$65 000 to C$200 000 per hectare are expected just in fisheries production directly related to the wetlands over the next 20 years.

U.S. wetland mitigation and restoration activities at the federal level restored or enhanced 526 000 hectares of wetlands with fiscal year 1988-1992 funds. Funding for these activities has increased from US$600 million in 1993 to US$800 million in 1995 (Zinn and Copeland 1996). In response to the 1993 and 1995 floods in the U.S. Midwest, the federal government offered to acquire easements on lands subject to severe flooding along the Mississippi and Missouri Rivers. These easements will provide flood buffer capacity, will help to restore lost and degraded wetlands and will protect fish and wildlife resources. Over 120 000 ha of land were enrolled in the program as of January 1996, and farmer interest far exceeds available funding (Zinn and Copeland 1996).

A 160 km section of the Kissimmee River in Florida is being restored to its pre-development state. Starting in 1954, the development relocated the river into an 85 km long channel at a cost of US$32 million. Restoration work will extend well into the next century and will cost US$422 million (Hardin 1993, Kirby 1993).

Since 1985, nearly 15 million ha of marginal farm land (8% of the total U.S. agricultural land area) have been retired from farming operations, reducing soil erosion and water pollution, and improving wildlife habitat. At a net program costs of US$8.9 billion, the benefits include reduction of 5 billion to 6 billion tonnes of top soil erosion in the first 11 years of the program, and a minimum of US$13.4 billion in wildlife-related, water and air quality, and soil productivity benefits. Recreational benefits, improved fisheries, lower dredging expenses, enhanced real estate values of riparian properties, and reduced flooding and disaster payments were not counted. Agriculture-related employment declined marginally and was probably offset by fish and wildlife-related jobs. (AFS 1996). The 1993 Midwest flood alone cost the U.S. taxpayer about US$15 billion in post-disaster relief, much of which was lost two years later when the smaller flood of 1995 occurred (Rasmussen 1996).

In British Columbia, the 1993 decision by the government to create the Tatshenshini Park involved trade-offs between conservation values and an estimated C$6 billion worth of copper ore deposits located in the park area.

Decisions to restore and conserve wetlands, tracts of forest, farmlands and urban parklands from highly profitable development indicate that society’s willingness to pay for the full external environmental benefits can be valued at tens or hundreds of thousands of dollars per hectare per year.

To avoid double counting, the costs of water pollution and land use not associated with the direct extraction, transportation, processing and refining of materials and energy products have been considered in Sections 4.3, 4.5 and 4.6.
 

4.4.1 Energy Consumption

Energy consumption must account for the impacts from exploration, extraction, transportation, processing and distribution of fossil fuels. A number of analysts have attempted to estimate the external costs of energy consumption, usually to determine the benefits of energy conservation. We have found no comprehensive estimates of the environmental costs of petroleum production (excluding social and economic costs). Much of the commentary on this issue focuses on damage from accidental oil spills during transportation. This omits other environmental damages, such as land use impacts of oil exploration, extraction and processing, and releases from oil wells and at processing facilities.

An average cost for cleaning up large spills can be derived, but it is important to note that these clean-up efforts are not fully effective, and cost estimates often focus on market costs and give little weight to non-market values. In addition, small and unrecorded spills are often not included in estimates of total oil pollution damage costs. One estimate provided by Lee (1993) values the annual environmental damages of spilled oil in the United States at $3.3 billion (1993 Canadian dollars), which he states is a guess. The value considers both major and minor spills and is derived from cleanup costs and some combination of improved navigation aids, vessel traffic services, additional hull structure (double hull, double bottom, mid-deck), human factors and improved mitigation.

Chernick and Caverhill (1989) estimate average petroleum marine oil spill costs by multiplying the minimum Exxon Valdez cleanup cost estimate of $1.28 billion (1989 US dollars) times 5 (because the cleanup only collected about 20% of total oil released), for an average cost of $202 per litre spilled. They consider this estimate conservative. Obviously, ecological damage inflicted prior to the cleanup and residual damage from the uncollected oil, are not counted. Hence the value is a lower bound.

Based on Chernick and Caverhill (1989) estimate, $200 multiplied by an estimated 10 million litres of oil dumped and spilled into the straits of Georgia and Juan de Fuca per year (BCRTEE 1992), an oil-spill portion of the total cost of resource use would be $2 billion per year. Not all of the oil shipped through B.C. waterways is for end use in the province. If it was, B.C. transportation sector’s share would be about $680 million ($2 billion times 34%, the latter being the share of transportation in end-use CO₂ emissions from fossil fuel combustion (British Columbia 1995)). About 70% of transportation fuel is used by road vehicles in B.C., based on the share in end-use CO₂ emissions (Ward 1994). Consequently, B.C. road vehicles would be responsible for about $480 million of oil spill costs in marine waterways alone. The cost of spills in inland waters, on land and underground is not included.

As water pollution is only one component of energy consumption costs, this value greatly underestimates the total cost. Other environmental damages from exploration, extraction, transportation, processing and distribution of fossil fuels are not accounted for and should be included as non-monetized costs in any complete analysis.
 

4.4.2 Vehicle Manufacturing and Road Construction

The materials considered the most significant by weight in vehicle manufacture and road construction are: steel, aluminum, glass, iron, plastics, zinc, aggregate, copper, bitumen and concrete.

During 1992, there were almost 120 000 first-time passenger vehicle registrations in British Columbia (R.L. Polk & Co.). This value approximates the number of purchased vehicles and together with the values in Table 4.15 it can be used to calculate the average resources consumed annually for passenger vehicle manufacture, allocated to B.C. residents.

Tables 4.16 and 4.17 are examples of spreadsheets that can be used to quantify the cost to the environment of materials used for vehicle manufacture. Similar tables can be created for roadway facility construction. Currently, cost calculations remain incomplete because of limited estimates of unit emissions for energy production and material processing. Further research is required to provide the full range of specific estimates for vehicle manufacture and roadway construction, and for costs that are more difficult to quantify, such as land degradation through mining, and option value for future generations.

DeLuchi (1991) has estimated the emissions of greenhouse gases from the full life cycle of light and heavy duty motor vehicles and the fuels they use. The United States Environmental Protection Agency (USEPA 1991) has estimated the emissions from the manufacturing of various materials such as steel and glass. Further research is needed so that our spreadsheets can include the full range of identified materials and unit costs.
 

Total Cost

The marine oil-spill portion of the total cost of resource use in British Columbia could be $2 billion per year, of which $480 million would be road motor vehicle share. The impacts of emissions, land pollution and forgone future use of petroleum products as well as the costs of other building materials and other resources and associated emissions are not accounted for, and should be included as non-monetized costs in any complete analysis. Although work toward costing vehicle manufacturing and road construction impacts is progressing, they will also have to be treated as non-monetized costs until further data is available. As the information necessary to account for the total range of impacts becomes available, it will become possible to monetize the cost to the environment of resource use. In the meantime, the external costs of resource and energy use should be considered qualitatively.
 

To avoid double-counting, only those costs associated with wastes that result from maintenance and replacement of vehicles and transportation infrastructure are considered in this section. This will prevent confusion when accounting for the environmental costs of resource production, water, land and air pollution resulting from oil spills, oil and materials production, and other sources of emissions and effluents not directly resulting from intentional disposal.

Waste Fluids and Scrapped Vehicles

While Johnson (1995a) has calculated the amounts of waste fluids, batteries, tires and vehicles scrapped in British Columbia, little research has been done to monetize the environmental costs of waste generated by transportation. A study by Lee (1993) estimates external disposal costs for auto bodies and waste oil. While these are based on proxy values rather than actual damage cost estimates they can be used to indicate the lower bound environmental costs of several components of transportation-generated waste. For the purposes of this report, the value provided by Lee (1993) for waste oil has been adopted for all engine fluids (Table 4.18). The environmental costs of lead acid batteries and tires appear to have been internalized through successful user-funded recycling programs so no cost estimates are provided for these items.

On a per capita basis the 313 ha of land used by wrecking yards equates to approximately 1 ha per 11 000 British Columbia residents. Although this amount is relatively small, these sites are usually close to urban areas and impose environmental and aesthetic costs. Poorly managed wrecking yards discharge pollutants into water supplies and contaminate local soils. Therefore, the environmental costs of wrecking yards should not be ignored.

Construction and Road Wastes

While the environmental impacts of construction wastes cannot be completely discounted, it appears that most environmental costs have been eliminated in British Columbia through progressive recycling programs and technological solutions. Certain areas in the province, such as the Victoria Capital Regional District, have even banned construction wastes from land fills, further encouraging waste reduction and recycling programs. Recycling of asphalt and sub-base products has proven, in most instances, to be more cost effective in British Columbia than the alternative of producing new road surfaces and discarding the pre-existing materials. The recycling of cement and asphalt materials has not only reduced the environmental costs of disposal, but also of resource use and air pollution. Cost savings act as the incentive for recycling waste concrete products. For example, a highway construction project in Wyoming incurred cost savings of up to $161 000 per kilometre by using recycled concrete as roadbed materials rather than crushed rock (Lauritzen 1991). Demolition rubble is routinely used in the Netherlands as road base material.

Estimate of Total Waste Disposal Cost

Although their magnitude is difficult to calculate, transportation waste disposal imposes small, but not-insignificant external costs. Some costs that were significant in the past are being internalized through recycling and disposal programs that are funded by user fees. However, these programs do not capture all wastes, particularly used oil and other fluids produced by home mechanics. Based on the unit values provided by Lee (1993) the estimated total cost is $10.1 million for 1992 waste predictions for the province of British Columbia. Averaged over 22 billion annual kilometres travelled by vehicles in British Columbia the cost of waste disposal is $0.0005 per km. This cost is a lower bound estimate, as it does not include wrecking yard impacts, for example.

It is difficult to place a dollar value on water quality and flow. Even after the quantities of pollutants and their impacts are known, the problem of valuing ecological functions exists. Economic costs for stormwater facilities, siltation removal and drinking water supply may be easier to calculate, although they are imprecise and incomplete indicators of total environmental values of water. Currently no standard economic analysis models identify the ecological and economic costs resulting from additional pavement areas, increased driving or other transportation activities.

New laws and policies designed to reduce pollution, prevent underground fuel tank leaks and internalize cleanup expenses may reduce some impacts, so it could be argued that current motor vehicle use imposes lower costs than has occurred in the past. Although these pollution control and management efforts may reduce these impacts, past damage and residual impacts are substantial. Growth in population density and total driving and public concern about water quality may increase the total costs of these impacts even if impacts per automobile kilometre decrease.
 
 

4.6.1 Review of Existing Estimates

A few studies attempt to measure water pollution costs (Table 4.19). Most studies focus on just one or two impacts, incorporating only part of the total cost. Even Moffet and Miller (1993) mention only a few of the total number of water quality and hydrologic impacts in their calculation of this cost. The Washington State Department of Transportation estimates costs of compliance with flood and pollution control requirements for state highways (ENTRANCO 1992). A few studies have examined the environmental costs of road de-icing, as shown in Table 4.20 (Field and O’Shea 1992).

Information from a number of studies can be used to estimate the total water pollution costs from roads and motor vehicles. State highways account for approximately 5% of US road length, or 10% of lane kilometres, and account for approximately 50% of US motor vehicle travel (US DOT 1992). An estimated 105 million off-street commercial parking spaces add approximately 10% to total road surface area or about 50% to urban road surface (Lee 1993). Total water pollution and hydrologic impacts are significantly greater than just state highway impacts, although exactly how much is uncertain.

State highway runoff impacts are conservatively estimated here to represent one-third of total roadway and parking surface runoff impacts. Using a simplifying assumption that Washington State highway water quality and flood control requirements (ENTRANCO 1992) represent runoff mitigation costs, the middle value estimated cost of meeting Washington State’s highway runoff mitigation requirements ((US$75 million + US$220 million) / 2 = US$147.5 million per year) is tripled to include non-highway roads, parking spaces and residual impacts (US$147.5 million x 3 = US$442.5 million). Using a simplifying assumption that every state has identical runoff mitigation requirements, this value can be scaled to the entire United States road system (US$442.5 million x 50) for a total annual national runoff cost of US$22.1 billion. Murray and Ulrich’s estimate of road salting costs (US$4.7 billion) can be added, for a total of US$28.8 billion per year.

This total could be distributed over the total number of road surface miles in the United States road system, or over the total number of vehicle miles travelled, depending on the requirements of the study being undertaken. Since water pollution increases with traffic, one can make the simplifying assumption that both increase in the same proportions, and assign total costs to total vehicle miles travelled. This total averages US$0.013 per automobile mile or about C$0.01 per vehicle kilometre. Note that this estimate does not include costs of residual runoff impacts, shoreline damage, leaking underground storage tanks, reduced groundwater recharge due to pavement, increased flooding or reduced wetlands, so it can be considered to underestimate total water quality and hydrologic costs.

According to Waste Management Group (1992), stormwater hydrologic impacts can be as damaging to the environment as pollution loads, implying that the cost estimate described here may need to be doubled to $0.02 per kilometre travelled to incorporate additional hydrologic costs. Thus, the estimates developed here can be considered an underestimate of total water quality and hydrologic costs. However, since this simplified model does not take account of the fact that de-icing or flood control costs are more directly correlated to road mileage than to traffic volumes, the estimate should be used with caution.

Further research will be needed to better define water pollution and hydrologic costs specifically for British Columbia and different vehicle types, and to determine whether other factors should be included in cost estimates. Potentially the largest component is the long-term impact of water pollution on ecosystems. Precautionary principle should be considered in re-estimates of water pollution costs.

The roadway investment evaluation models of both the Swedish National Road Administration (1986) and the Danish Road Directorate (1992) incorporate methods for quantifying barrier effects on specific lengths of roadway. These methods involve two steps. First, a barrier factor is calculated based on traffic volumes, average speed, share of trucks, number of pedestrian crossings and length of roadway under study. Second, the demand for crossing is calculated (assuming no barrier existed) based on residential, commercial, recreation and municipal destinations within walking and bicycling distance of the road. The Swedish model also adjusts the number of anticipated trips based on specific traffic, population characteristics and land-use conditions. The Scandinavian models can be used to develop estimates of barrier costs for typical North American road and traffic conditions.

Gylvar and Steen (1983) estimate that the barrier effect represents 15% of roadway costs considered in Dutch cost benefit analysis. Total costs are 50% economic (travel time, accident reduction and vehicle operating costs), 30% noise, 15% barrier effect and 5% air pollution. A recent re-evaluation of the proposed British motorway network expansion revealed that a substantial cost item due to barrier effects was missed in previous cost-benefit appraisals.

Freund and Martin (1993) quote an empirical study of decreased mobility of British children due to increased street traffic. The increase in the personal freedom and choice arising from widening car ownership has been gained at the cost of freedom and choice of children. The result has been increased parental escort of school children, especially in autos, at a considerable increase in cost of between £10 billion and £20 billion in 1990.

Sælensminde’s (1992) data makes it possible to relate barrier costs to noise costs in Norwegian urban areas. The total costs of barrier effects are about equal to the estimated total costs of damage due to noise. Based on this information, we estimate the shadow price of barrier effects to be of the same order as the shadow price for traffic noise costs to a community (C$1000 to C$1500 per person affected per year). The costs of wildlife habitat severance and farming community severance have not been estimated yet. Most of these costs will probably remain intangible.
 

The word biodiversity is used here to refer generally to the range, complexity and abundance of species, genetic diversity and ecosystems that make up the biosphere. It is possible to identify the many ways in which humans value biodiversity, as shown in Table 4.21, but not to put a finite value on the whole.

In theory, it is possible to price various components of biodiversity such as individual species and ecosystem services. In practice, to do so is extremely difficult. Many ethical, social and ecological questions need to be answered before a representative dollar value can be created that may be used in any form of valid analysis. These dollar figures would have to fully account for the consumptive, non-consumptive, existence and option values.

Both species and ecosystems have value and are intertwined components of biodiversity. Interacting species compose ecosystems; therefore, the value of an ecosystem could be considered the sum total of its species. However, a simple additive framework is unlikely to be sufficient, since it fails to take account of synergistic effects and interrelationships. The value of an ecosystem almost certainly exceeds the species aggregate. Table 4.21 illustrates an additive calculation framework. Of course, the simple fact that species compose ecosystems cannot be forgotten.

4.8.1 Biodiversity and Monetization

The most common strategy used to estimate biodiversity is to attempt to sum the values of its individual components (Norton 1988). This method assesses individual species on the basis of their values, as suggested in Table 4.21. It is unlikely that this approach will ever be fully successful. To account for the full economic value of biodiversity we would need to identify all species and their uses, place a dollar value on those uses and estimate the likelihood of any additional uses so that the option value could be accounted for. Unfortunately, the total number of species that exist on the planet is unknown. Estimates place the range from 3 to 30 million (May 1992). At present, only a small percentage of species has been identified, and we know very little about most of those species. To apply a dollar estimate to each of the potentially 30 million species is unrealistic.

A second problem results from synergistic effects. Species do not exist independently. Most ecosystems are webs of interdependent life. A species may depend on just one other species for food or it may be involved in a complex of interrelated species. Species on which others depend then have value to those species. Furthermore, the extinction of one species may directly lead to the extinction of others. As ecologistsare only beginning to understand the fundamental relationships of simple ecosystems, the possibility of understanding the most complex, such as tropical rainforests, appears remote.

The valuation of biodiversity also involves ethical considerations. There is a danger that one species may be considered economically expendable simply because it has a relatively low individual value to humans. Another problem is that if a complete and definitive dollar value could be found for a species, humans may treat that species as expendable within an analysis framework based on its dollar value. If one elk is worth exactly $1 million and a highway generates $5 million in benefits, are five elk expendable in a trade-off of elk versus human utility generated from the highway? The question remains, whether humans have the right to determine the existence of individuals of other species simply based on economic criteria, even if all the values of species and ecosystems were monetizable.

Clark (1973) provided an example of moral objections exceeding economic benefits.From an economic standpoint, it would be advantageous to harvest every blue whale in the ocean as quickly as possible and invest the earnings where they could gain a greater rate of return. This was based on the observation that blue whales have a reproductive rate of 5% to 10% annually. If the whaling fleets used an annual discount rate of 20%, the most profitable route would be to exterminate the whales and reinvest the capital.

It is likely that global public outcry, representing existence or moral value, would outweigh the simple economics of harvesting and prevent whaling to extinction. Highly visible species, such as whales or pandas, appear to have a very large existence value in that people would be willing to pay almost anything to preserve them. It remains very difficult to quantify, monetize and use existence values within an analysis framework. Beyond this, humans fail to assign existence value to unattractive or unknown species. Whether usefulness or attractiveness to humans is an ethical basis for valuation remains debatable. These are questions that cannot be answered easily, but must be considered when attempting to put a definitive dollar value on biodiversity.
 

4.8.2 Current Estimates

Given that it is unlikely that a complete accounting of all the potential values of the components of biodiversity will or can be calculated, it is clear that wildlife and forests have significant economic value. For example, Environment Canada has estimated that in 1987, $1.06 billion was spent on recreational hunting and $4.04 billion on non-consumptive and other wildlife activities. This same survey revealed that B .C. led the country with the highest per capita expenditure on wildlife-related activities at $355.

Reid (1990) estimated that between 1988 and 1989 B.C. residents and non-residents spent $144.3 million on hunting activities, residents spent $611.1 million on non-hunting activities and $9.2 million was spent on trapping expenses. Residents indicated that they would be willing to contribute nearly $131.8 million voluntarily to preserve the current abundance and variety of wildlife. Stone (1988), estimated that the net economic value of resident fresh water angling in British Columbia was $89.1 million in 1985. The total value of fish and shellfish caught in British Columbia in 1991 was $363 million.

As a consequence of productive forests, the forest industry directly generated more than $710 million in 1992/93 from stumpage, royalties, fees and taxes and provided nearly 72 000 direct jobs in 1991 (MoF 1994). In the United States, an estimated US$24 billion is spent annually on sport fishing, which generates 900 000 jobs directly and US$70 billion to the total economy when multiplier effects are included (AFS 1996).

Table 4.22 provides additional estimates that illustrate a portion of the economic value for selected species and ecosystems. Total monetary values of various habitats are substantial, but so are the total recreational, option, existence and bequest values of individual species and wilderness when added up over all households or individuals in the country. Ecological economic values include the value of recreational activities, but does not include the very substantial contribution of option, existence and bequest values.

4.8.3 Estimate of Shadow Prices for Biodiversity

As a very rough estimate of the order of magnitude involved in valuing biodiversity, one can consider the estimates in Section 4.3 of the economic value of lost or impaired ecological functions and degraded habitat. The estimate can be regarded as a lower bound on the economic value of biodiversity. According to Table 3.27, Canada, through various governments, currently preserves 70 million hectares for the sake of biodiversity and associated values. This represents 7% of Canada’s total area. Based on the values provided in Section 4.3.3, it can be assumed that about $10 000 per hectare in annual economic benefits is available from an average habitat. The annual worth of these Canadian reserves would be at least $700 billion, or about the order of Canada’s GDP. The 12% of the total land area of Canada targeted by the United Nations convention on biodiversity would be worth at least $1.2 trillion per year, and the 24% recommended by scientists would be worth at least $2.4 trillion per year.

These figures perhaps convey a sense of the economic utility of biodiversity. Unknown and unpriced functions and values of biodiversity would be extra. While these figures remain too general to use as a reasonable shadow price, they should be kept in mind. It is likely that most values of biodiversity will remain unmonetized for the foreseeable future, but their overall magnitude, even from a restricted economic perspective, should be appreciated when ecological values are considered in a non-monetized framework