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Economic Instruments for Managing Water Quality in New Zealand

Executive summary

Acknowledgements

1 Introduction

2 Overview

3 Developing a framework for evaluating economic instruments

4 The role for non-market valuation

5 Economic instruments for diffuse source water quality issues in New Zealand

6 Summary and conclusions

References

2 Overview

Economic instruments are a broad class of mechanisms that seek to align private incentives with public natural resource management (NRM) goals. Expectations regarding what can be achieved using economic instruments should be dependent on how well the instrument has been tailored for the circumstances in which it is to be applied. Ideally, economic instruments will:

• Achieve a given level of environmental improvement compared with a prescriptive regulatory approach for a lower cost; or
• Create markets for environmental outcomes.

Inclusion of economic instruments in the policies used for NRM is underway or being considered in many OECD countries including New Zealand, Australia, Canada and the United States (Stavins, 2000). The approaches being used or proposed fall into several distinct categories. Most instruments involve modifications to existing "non market-based" policy to include market mechanisms. One example is using a tendering approach or an auction process to distribute public funds. Another example is tradeable permits which involve modifying regulation to allow for buying and selling permits.

2.1 What are economic instruments?

A series of names are currently being used in the economic and policy literature to describe economic instruments (Hahn 1989; Sterner 2003; Randall 2003) and these include:

• economic instruments
• market-based instruments (MBIs)
• incentive-based policy
• policy instruments
• flexible instruments.

The term, market-based instruments (MBIs), has become the standard terminology in countries such as Australia. Changing the language and abandoning the term "economic instruments" was important in shifting the policy debate from an approach advocated by economists and business people to an approach that included a wider circle such as natural resource managers, planners and regulators.

A typology for categorising economic instruments is presented in Text Box 1. Economic instruments rely on a well-defined regulatory framework, and in some cases property rights and entitlement systems, in order to send market based signals to change behaviour for the benefit of the environment.

At the outset it is important to emphasise that economic instruments are not a "free market" substitute for all NRM policy approaches such as command and control, but rather economic instruments involve modifying traditional policy approaches to include market mechanisms. The need to investigate and set goals and targets for environmental quality either remains the same or increases. Similarly, the need for research to understand the biophysical processes remains the same or may increase for NRM agencies that wish to utilise economic instruments.

An attractive feature of economic instruments in NRM policy is that the approach can align private incentives for profit with public goals of improving environmental quality. Often individuals and firms have detailed knowledge of the opportunities that exist for them to improve their interaction with their environment compared with a regulator. Policies with incentives related to NRM outcomes reward individuals who use this detailed and specialised knowledge to reduce the environmental impacts of their activities in innovative, low cost ways. For example, an irrigator facing a charge related to water use may be motivated to implement low cost ways of reducing water use. It is important to acknowledge that there are some significant constraints limiting when economic instruments can be effectively applied. For example, some economic instruments require the development or refinement of property rights, monitoring capacity, reporting and accounting systems. If the initial investment required to develop such systems is large, or the administrative cost in running them is large, the benefits of these economic instruments may not be sufficient to offset their costs.

Some economic instruments require changes to property right regimes that may not be politically feasible, implying that it is not possible to utilise certain economic instruments. The attitudes of individuals and businesses toward some of the economic instruments available may preclude their use in some catchments. In any of these situations, it may be necessary to choose from a reduced set of economic instruments and/or rely on traditional command and control approaches to environmental management. It should be noted that what is perceived as politically acceptable evolves over time and can be influenced through policy designed to inform and influence.

The remainder of this section outlines the major types of economic instruments that will be considered for use in managing diffuse source environmental water quality issues.

A basic description of each type of instrument is provided and examples of applications that are particularly relevant are outlined. The discussion focuses on:

• environmental charges
• tendering
• compensated covenants
• land leasing
• transferable (or tradeable) permits
• environmental offsets.

Environmental charges, tendering, compensated covenants, land leasing are all incentives that affect the cost of using or conserving a resource and are price-based incentives. Transferable permits and environmental offsets affect the overall quantity of the resource and are referred to as quantity-based incentives.

2.2 Environmental charges

Environmental charges aim to reduce the level of a discharge to the environment by charging a fee per unit emitted. A closely related approach involves reducing the amount of some tax or fee that an individual is required to pay in exchange for undertaking some activity that benefits the environment. An example is making expenditures on solar hot water heater tax deductible to address green house gas emissions. Both environmental charges and the related tax or fee reduction approach create private incentives to reduce environmental impact by offering opportunities to minimise environmental charge payments (or maximise tax rate reductions). In some settings, charges can induce innovation and result in declining discharges over time.

The choice of whether to use charges or tax rate reductions depends on social and political values about whether the polluter should pay or the beneficiary (society) should pay for the environmental improvements. Charges are clearly an application of the polluter pays principle, while tax rate reductions are an application of the beneficiary pays principle. The person that takes that action with positive environmental impact receives a public payment in the form of a reduced tax burden.

However, as environmental quality has deteriorated or come under threat in New Zealand, there has been a trend towards introducing fees charged to those whose activities adversely impact the environment. This is broadly reflective of trends elsewhere. In the United States, Australia and the EU, there has been a distinct move towards setting charges in proportion to the quantity of emissions. Some European countries such as Denmark and the Czech Republic, have been implementing charges targeted to municipal wastewater dischargers and industry. In the Netherlands, there are significant charges for nitrogen and phosphorous use on farms in excess of set limits. Australia as part of its National Competition Policy is reviewing how environmental externalities are incorporated in urban and rural water charges (National Competition Council, 2003).

Charges can be designed in any number of ways and for any number of reasons including raising revenue. If the purpose is to send a clear signal to the polluter, then the charge would require each polluter to pay a charge equal to the cost of the individual incremental damage associated with their discharge. Economists refer to this ideal charge as a Pigovian tax. Pigovian taxes result in each polluter having the incentive to reduce discharge until the cost of avoiding the last increment is equal to the value of the last increment of damage avoided. Actually implementing this type of ideal charge in the real world is simply so difficult that it is rarely, if ever, done. The actual physical damage per unit discharge is often not understood with any certainty and the value of the damage done is equally difficult to determine. Even attaining approximate values often involves considerable and expensive research. The "Pigovian prescription" also involves charging each polluter differently. Attaining such fine differentiation involves a level of administrative effort that is typically not feasible or cost effective. In addition, such differentiation is likely to be seen as inequitable by many and can be subject to legal challenge.

The result is that most charges are "second best" approximations to Pigovian charges given real world information and administrative cost limits. For instance, a charge that retains some properties of a Pigovian tax is a per unit charge that increases in "tiers or blocks" reflecting the nature of the damages. The NSW EPA Load-based Licensing programme for regulating industrial facility pollution, described in Text Box 2 (from Bright et al, 2002), is a good example of a charges approach. The programme involves a "tiered pricing scheme". Charge rates increase when discharges exceed prescribed limits. This reflects the increasing cost of higher rates of discharge to the environment and other users. An alternative second best charging mechanism, where the impact of discharging pollutants is not uniform across location, is zone-based charges.

Charges may not necessarily represent the best policy choice in situations when discharges over a certain threshold level cause significantly more damage than discharges below the threshold level. This is because the level of discharge that can be expected is determined by responsiveness to price signals that may not be well understood before charges are implemented. Responsiveness to price will also change over time with factors like technology. If the response to charges have been underestimated, [That is, a much larger increase in price would be required to reduce the demand for the input associated with the pollution.] much greater than anticipated environmental damage can result from charges. Quantity-based approaches, in contrast, involve a discharge rate and thus ensure a particular outcome.

A charging approach is most easily applied to situations where charges can be set in proportion to the environmental outcomes of interest. The prerequisite is that outcomes can be monitored in a relatively straightforward manner. When monitoring is difficult and expensive, implementing charges will typically require significant scientific effort to relate outcomes to levels of some easily measured "proxy". A proxy is an input or outcome which is correlated with the outcome of interest. For example, the level of manure application to fields might be a proxy for nutrient loading from manure application. The benefits of implementing charges can become more difficult to justify when charges are not based on direct outcomes. This is because the costs of establishing a relationship between the proxy and outcomes can be expensive. Typically, outcomes are less than perfectly correlated with proxy input use. In practice, the use of such charges is much easier if it is made clear that the aim is to signal the general nature of the cost rather than to set the charge with precision.

2.3 Tendering, compensated covenants and land leasing

2.3.1 Tendering

Tendering is being used by NRM agencies to distribute public funds across private firms and individuals who engage in activities to improve the environment. Tenders work by having those interested in the public funds compete by submitting an offer to undertake work for a price. The tenders describe the on-ground works that will be provided, an assessment of the likely benefit and the payment that would be required to undertake the works. The NRM agency then ranks all tenders received on the basis of cost per unit of environmental benefit offered. Tenders are accepted in order of cost per unit of environmental benefit offered until the budget is exhausted.

With this tendering system, each individual, rather than the government agency, sets the cost share rate they would be willing to accept. The tendering approach allows landholders who are willing to undertake environmental improvements at low cost-sharing rates to do so. The result is that a higher level of environmental improvement effort can be attained for a given expenditure than is achievable when a single cost sharing rate is offered to all. Text Box 4 below from Connor and Bright (2003) describes the Victorian DNRE BushTender approach to allocation of biodiversity conservation cost sharing. Text Box 5 presents an example of tendering already in place in New Zealand.

Recent work by Hailu and Schilizzi (2003), using an agent based modelling approach, suggests that the efficiency gains from BushTender may be limited to the first few years. After the first few years, the least cost bidders have completed their on-ground works and the "expensive" bidders have dropped out. The remaining bidders tend to provide bids that converge towards a band of acceptable bids. Applying this concept to water quality issues, the early rounds of tendering allows the NRM agencies to identify the landowners or firms that are the most enthusiastic and tendering allows NRM agencies "to pick the low hanging apples". Sustained cost savings therefore may not occur over successive rounds as landholders "learn".

2.3.2 Compensated covenants

As the use of economic instruments has increased, an issue has been emerging with respect to how governments can protect publicly funded on-ground works on private land. More generally, the question is one of how can governments ensure the ongoing provision of environmental benefits/ecosystem services on private land after a payment has been made. These two inter-related questions can be addressed through different types of agreements, easements and covenants that essentially contract a land holder (and sometimes their successors) to continue providing the benefits in exchange for tax concessions or nominal payments.

Governments have been experimenting with ways to protect the public interest on private land. The approach can be as general as a management agreement where the landholder and government enter a contract regarding how the land is to be managed for a specified time period and payment schedule. Examples of this include the Land Retirement Programmes, Farm and Ranch Lands Protection Programmes, and the Debt for Nature programme where development rights are transferred in exchange for the cancellation of loans with the Farm Service Agency in the USA. Examples from the UK include the Countryside Stewardside Scheme where landholders enter into 10-year agreements to manage land for habitat in return for annual payments. In South Australia, once-off payments have been used to negotiate covenants, known as Heritage Agreements, in perpetuity.

The basis for most programmes has been a contractual arrangement where the landholder agrees to undertake some activity or refrain from some activity in return for some financial incentive such as a payment, tax concession or cancellation of loans, etc. Contractual arrangements provide a high degree of variability and can be tailored to the situation such that the agreement provides the legal protection that parallels the conservation value (Binning and Young, 1997).

In New Zealand, covenants are used to protect important ecological assets. Examples include the blocks of forest in the Eastern Bay of Plenty and the Aorangi-Awarua project. Maori landholders can protect indigenous ecosystems under a Nga Whenua Rahui kawenta. These covenants provide long-term protection with provisions for inter-generational reviews of the agreements. Funding for fencing, pest control and other management costs may be provided through the Nga Whenua Rahui Fund.

2.3.3 Protective land leasing

In New Zealand, NRM agencies will purchase a leasehold interest in privately owned land in order to directly undertake land use practices that are necessary to achieve particular protection or conservation outcomes. An example is Kapenga Wetland which is being leased from Maori Farm Trustees.

Opportunities also exist for using other forms of land leasing where public land is leased for agriculture on the condition that best management practices for water quality are implemented. One condition might be to require that environmental management systems that focus on aspects of erosion control features be developed and implemented.

2.4 Transferable permits

Transferable permits involve setting an initial overall limit (cap) on the emissions or pollutants. A system is then set up for allocating the amount that a group or individuals are allowed to emit or to pollute in the form of a permit. Individuals are only allowed to exceed their initial allocation if they purchase additional permits from someone else. Those who can achieve large reductions in pollutants at low cost are motivated to sell part of their allocation of permits at a profit to others. The approach encourages producers to think about the external impacts of their production activities and to search for innovative, low cost ways to reduce them. As a result, transferable permits can allow achievement of a desired level of emission at considerably less cost than uniform environmental standards can.

Transferable permit approaches tend to be preferred to environmental charges by producers who anticipate that they may be able to profit by producing significant reductions in emissions. In contrast, with a discharge fee policy, all producers pay the environmental agency, unless they produce absolutely no emissions. In effect, most permit systems use a market mechanism to set the lowest charge needed to achieve an outcome and then redistribute the resultant revenue amongst producers. In contrast, under a charging system, the resultant revenue is usually kept by the agency.

Experience with transferable permits suggests that establishing rules for trading is not sufficient to guarantee trades will occur. The Fox River in Wisconsin is an example of a programme where very few trades have occurred. At best only a thin market was created due to the restrictions on trades designed to protect third parties from local adverse water quality impacts of trade and the uncertainty created for the polluters concerning renewals and how trade might affect renewal rights. An important implication is that when significant restrictions on trade are necessary to protect third parties, a tradeable permit approach might not be a good instrument choice.

Ex-post evaluations of programmes in the USA, suggest that for point source pollution problems, transferable permits have had some success. The Acid Rain programme was successful in demonstrating that large scale trading can occur and well-designed programmes can be resilient to unexpected changes or surprises. When compared with command and control regulation, it is likely that transferable permit systems were better at adjusting to sudden changes such as the deregulation of the railway systems which allowed new sources of low sulphur coal and significant fuel-switching to occur. Smith (2002) suggested that design mistakes to be avoided include allowing too much flexibility or bonus allowances in the early phases to ease the adjustment process.

Transferable permit programmes are being considered and in some cases implemented in many forms across the industrialised world in the area of green house gas emissions, renewable energy, solid waste and other non-point source pollutants. While Raux (2002) suggested that while transferable permits are particularly challenging for diffuse source pollutants, the approach has been used for such problems with some success. An interesting example is presented in Text Box 6 for on-farm nutrients.

Finally, water quality is often related to the quantity of water being pumped whether it is a groundwater or surface water system. Young and Hatton MacDonald (2003) have suggested that water quality issues can be accounted for in a groundwater trading system by finely differentiating rights to water. Specifically, the right to water is separated into a right to receive a periodic allocation of water and a right to use the water. This allows for trading possibilities with respect to the water, salinity and other water quality issues.

2.5 Environmental offsets

Environmental offsets are most often used where environmental quality goals are only just being achieved, or where there is non-attainment, and there is pressure to allow further development. The idea behind offsets is to allow further development to occur, but in a way that either maintains or improves the stock of environmental capital. Thus it is possible to have further development that damages the environment, provided that works are undertaken to either offset or more than offset this damage. This is consistent with the concept of weak sustainability, whereby sustainability is achieved by maintaining the stock of environmental capital at some specified minimum level, possibly by "projects/ policies designed to produce environmental benefits" (Hanley, Shogren and White 1997, p.429).

In a catchment, for instance, developers may be required to implement best available technology (BAT) to control water pollution. However, development with BAT may still lead to a net increase in discharges. Under an offset programme, this increase would be offset by a reduction of an equal or greater amount somewhere else in the catchment. Developers either undertake the offsetting action themselves or pay for others to do it on their behalf.

Environmental offsets have been used in a range of circumstances. Broadly these can be grouped as involving either pollution or biodiversity/vegetation offsets. Examples of pollution offsets include air quality in the USA (as prescribed by the Clean Air Act), water quality offsets in the USA (see Text Box 7), and drinking water and air quality offsets in NSW, Australia [http://www.epa.gov.au/greenoffsets/index.htm]. For biodiversity/vegetation offsets the best known examples are wetland mitigation and streambank mitigation banking. This sort of environmental offset has grown rapidly in the USA. In 2001, there were estimated to be over 200 operational wetland mitigation banks in the USA, and over 100 awaiting regulatory approval (ELI, 2002). However, wetland mitigation banking is not without its critics. While it is apparent that wetland mitigation banking has led to much larger areas of wetland being preserved, there is evidence that the quality of the wetlands produced through offsets is inferior to that lost through development (Salzman and Ruhl, 2000). However, this is less likely to be an issue with pollution offsets.

When the development and offset actions have similar impacts - as is more often the case with pollution offsets - ensuring net environmental benefit from an offset policy is more straightforward. This is achieved by developing "trading ratios", which indicate the amount of offsetting required for a given development impact. As an example, a trading ratio of 2:3 implies that a development that introduces two tonnes of phosphorous into a waterway would require three tonnes of phosphorus elsewhere in the catchment to be offset. Trading ratios are usually based on several factors, including the desired environmental improvement, the level of uncertainty about environmental impacts, and the distance between the development and offset sites (Morrison, 2003).

Environmental offsets are usually undertaken prior to the development occurring so that there is no temporal loss in environmental quality, and to provide additional certainty that the offsets will be effective. In the case of water-based offsets (such as non-point source pollution), offset trades are also usually restricted to occur within a "geographical service area" so that no localised problems or hotspots occur.

Environmental offsets can be implemented via bilateral negotiations between stakeholders (i.e. where developers directly contract with owners of potential offset sites), or through privately or publicly owned offset banks. An offset bank is not a bank in the usual sense. Rather an offset bank involves the completion of one or more projects in which environmental remediation works are undertaken. By completing these works, offset banks earn "credits" which can then be sold to developers who are creating net-impacts on environmental quality (Morrison, 2003). In the case of wetland mitigation and streambank mitigation banking, offset banks are run by either private business, non-profit organisations or, in some states, by government organisations. The majority of wetland mitigation banks are privately operated (ELI, 2002).

Compared to other economic instruments, environmental offset schemes have a number of unique requirements. These include:

• development of enabling legislation (this is also required for trading schemes, but not for other instruments such as tendering where participation is voluntary)
• changes to licenses of new developments so that no net discharges are allowed
• the capacity to estimate changes in loads at development and offset sites (several software packages have been developed to do this), and
• development of offset banking instruments.

As a general rule, one of the main advantages of an offset programme is that they do not require all land users to be monitored. As a result they tend to be less expensive to implement than a tradeable permit system. Moreover, with an offset programme, there is no need to find a way to "allocate" permits in a manner that is politically acceptable. Other implementation issues for environmental offset schemes are discussed in Morrison (2003).

The use of environmental offsets is likely to be supported by developers because it provides a way of achieving further development in areas where there is no "unused" environmental capacity. Politically this may be a useful option where the benefits associated with development (e.g. jobs, regional income) may be high. In contrast, environmental offsets have received less support from green groups because of concerns that offsets have been of inferior quality to what is lost through development, and because of concerns that it will increase the rate of damage to the environment.

The economic theory behind the use of offsets has not been well developed (Baumol and Oates, 1989; Vernon and Goddard, 2003; Morrison, 2003). From an environmental management perspective, and from the perspective of the community, offsets are an appealing instrument because development can proceed provided that environmental damage is offset.

Efficient regulatory instruments should move an economy from a private market optimum to an optimum that maximises social welfare. An effluent charge (or Pigovian tax) does this by setting a tax equal to the damage cost of pollution. With any permit, offset or charging scheme, it is possible that the cost of the offsetting action (plus the cost of administering the scheme), may exceed the damage cost of the externality. This is likely to be the case in highly urbanised areas where the cost of offsets can be particularly high and require specialised engineering solutions and costly retrofits. In peri-urban or rural areas where there are more offsetting alternatives, this is likely to be less of an issue. As well, from a policy perspective the concern is not only with costs, but also with benefits and equity. An offset programme may ensure that the environmental capital stock is maintained, but a change in the composition of this capital stock may alter the benefits of the capital stock to the community, as well as affect the distribution of benefits. For instance, developing a wetland within an urban area and replacing it with a wetland some distance from the urban area could substantially reduce the local recreational value of the wetland area at one location but increase recreational opportunities at the location of the offsetting action. Overall, however, the net-recreational benefits to the community may fall. Further, a new wetland may not provide the same quality of ecosystem services that an existing wetland provides.

These warnings should not be seen as suggesting that offsets are likely to be an unworkable economic instrument. Cost-benefit analysis with careful statement of the basis for comparison will indicate the practicality of implementing an environmental offset scheme.

One important limitation of offsets is that they do not directly deal with existing pollution, except to the extent that developers more than offset their impacts. In this sense, offsets are not fully consistent with the polluter pays principle. New polluters pay to reduce their pollution, but existing polluters are not penalised. This characteristic of offset arrangement can sometimes be useful because it is harder to "claw back" existing entitlements than establish new standards, but it can also limit the effectiveness of the instrument when current entitlement are causing significant environmental harm.

Finally, it should be noted that there are opportunities to use environmental offsets together with other instruments. Offsets can be a good complement to other economic instruments such as trading schemes, traditional regulation, and environmental charges.