Action plan requirements can also be spelt out. A range of fertilizer applica- tion limits are set; for example, a maximum of 200 kg P2O5ha1can be applied on grassland. Application of animal manures is forbidden between 1 September and 1 February on soils vulnerable to leaching. Calculated farm manure sur- pluses will be taxed, based on compulsory construction of manure balance sheets for each farm. A manure trading system is also in operation, whereby farmers buy and sell the rights to spread manure, thus allowing farmers in ‘spare capacity’ areas to sell spreading rights to farmers in areas where no further appli- cations are allowed. These rights are defined on a comparison of actual livestock stocking rates on a given farm with maximum allowable spreading rates. The trading system has an inbuilt decreasing supply mechanism, whereby the entitle- ment is cut by 25% each time a trade occurs that is not tied to land. Finally, the government has supported industrial reprocessing plants for manure surpluses.
Helming (1997) produced some estimates of the implications of these policy changes on farm incomes and environmental quality in The Netherlands.
What lessons can be learnt from existing policy? And how can policy design be improved? Several themes seem to be important. Firstly, there is a need to improve coordination between agri-environmental policy and agricultural policy, and indeed other rural policies. Article 130R of the Maastricht Treaty calls for the greater integration of environmental and agricultural policies, yet we still find the CAP and agri-environmental schemes pulling in opposite directions. For instance, Environmentally Sensitive Areas in upland regions of the UK frequently suffer from over-grazing by sheep, which reduces their ecological quality. ESA payments thus encourage farmers to reduce grazing pressures by reducing sheep numbers, but the system of Hill Livestock Compensatory Allowances under the CAP encourages farmers to keep more sheep, by offering a headage payment per breeding ewe. This conflict, of which there are many other examples, both increases the apparent costs of conservation and reduces take-up rates.
Cross-compliance has often been suggested as a means of integrating envi- ronmental goals with the broader goals of farm policy. Many different forms of cross-compliance exist, such as ‘red ticket’ approaches, where farmers lose all entitlements to support payments if they fail to meet environmental standards,
‘green ticket’ approaches, where farmers get higher levels of support if they meet these standards, and ‘pink ticket’ approaches, where farmers lose part of their entitlements if they violate environmental conditions (Batie and Sappington, 1986; Baldock and Mitchell, 1995). The difficulties of implement- ing credible cross-compliance schemes are well known (see, for example, Spash and Falconer, 1997); however, there has been a limited but increasing use of cross-compliance in the EU (for example, environmental requirements for set- aside payments) and in Norway (where environmental conditions are attached to arable support payments). The UK government, for one, has called for the increasing use of environmental cross-compliance within the CAP in future, citing the example of environmental conditions being imposed on the arable area payment scheme, and on headage payments for livestock (DETR, 1997).
Secondly, there is a need for more targeting of agri-environmental programmes. Payments for the production of environmental goods from rural land management should ideally be targeted where these are most valued. For example, only 10% of the land area of Finland is arable land, and much of this is highly valued for its ecological and cultural values. Yet Finnish applica- tions of the Agri-Environmental Regulation (92/2078) have focused on nutrient run-off problems, rather than on conserving the biodiversity and landscape values of this land (Sumelius, 1997). In a different vein, efforts to protect habitats may be hampered because payment schemes are adminis- tered on an area-wide basis, rather than a habitat-specific basis.
Thirdly, uniform payment rates are inefficient if opportunity costs of meeting scheme requirements vary across farmers: there is ample evidence that this is indeed the case (Hanley et al., 1998a). Next, many voluntary pay- ment schemes are set up in terms of required management actions (e.g. a reduction in stocking rates, a reduction in fertilizer use), but links between management actions and environmental effects are often uncertain. Better
Agricultural Pollution Policy in the EU 161
then, some would argue, to target policy at environmental outcomes. As we know, this is a problem, due to the non-point nature of much agricultural pollution. Environmental outcome targeting may thus be better suited to policies aimed at the production of wildlife and landscape (since the identifi- cation problem is easier), whilst management action targeting is better suited to policies aimed at reducing environmental bads such as pollution.
Finally, it is apparent that very few environmental policies within agricul- ture in the EU are subject to any kind of economic efficiency criteria. These include cost-effectiveness and Pareto efficiency. With regard to the former, there has been little official recognition of the role that economic instruments can play in bringing about cost-effective solutions to externality problems.
Where economic instruments have been used, the main purpose has been either as a gesture towards the ‘polluter pays’ principle (e.g. nutrient taxes in Sweden and Finland), or as a means of improving target achievement (e.g. the Danish pesticides tax). To some degree, this is a function of the difficulties in designing taxes and permit markets for non-point pollutants, but mainly one suspects it arises from a failure on the part of economists to sell their case well enough. The direction of change is, however, promising, with The Netherlands planning to place more emphasis on the market and on taxes to achieve its responsibilities under the Nitrates Directive, and with the general use of economic instruments in environmental policy rising throughout the EU.
With regard to Pareto efficiency, economists would prefer to see more gov- ernments applying cost–benefit analysis criteria to the control of agricultural externalities. Whilst there is some use of this in the context of wildlife and landscape policy in the UK (as summarized in Hanleyet al., 1999), and whilst similar policies have also been appraised in this way in France, Sweden and Norway, there are a great many more instances where cost–benefit criteria are not applied. This may be leading to some big policy mistakes. Are the costs of meeting the Nitrates Directive in the EU really warranted by the expected environmental and health benefits? Should a much greater level of resources be put into agri-environmental schemes under regulation 92/2078? Are the costs of the Danish pesticides action plan too high? At the present, we just do not know the answer to these important questions.
Endnotes
1. BOD is biological oxygen demand, a common measure of the polluting potential of organic substances.
2. For a useful account of point source pollution control in Denmark, see Andersen (1999).
3. This is defined as a 50% reduction in the volume of active ingredients applied as well as a 50% reduction in the annual number of treatments per unit area (Dubgaard, 1999).
162 N. Hanley
The impact of international trade on the environment has been a contentious issue since the early 1990s. The debates over the North American Free Trade Agreement (NAFTA) and the Uruguay Round trade negotiations, which led to the creation of the World Trade Organization (WTO), stimulated a great deal of economic research on the ways in which international trade might be envi- ronmentally harmful or beneficial. Interest in the effects of trade on the envi- ronment then subsided until 1999, when protests in Seattle over a new round of WTO negotiations moved the issue to the forefront once again.
In theory, the environmental impacts of trade liberalization would not be a cause for concern if socially optimal, flexible and internationally coordin- ated environmental policies were in place in every country worldwide. In such a world, because environmental policies were flexible, any worsening of environmental externalities in some country due to trade liberalization would be mitigated by a socially optimal increase in the stringency and/or scope of that country’s environmental policies. Furthermore, any worsening of a global environmental externality such as biodiversity loss would be mitigated by an internationally coordinated and socially optimal increase in the strin- gency and/or scope of environmental policies worldwide.
Whether these conditions can be satisfied to a reasonable degree is ques- tionable, particularly in agriculture. One of the messages of the preceding chapters of this book is that designing cost-effective environmental policies that can adequately address environmental problems associated with agricul- tural production is a hard problem. The non-point character of agricultural pollution places constraints on the options available to policy-makers. One cannot control non-point pollution in the way that one can control point sources of pollution, such as the flow of sewage out of a pipe or pollutants
© CAB International 2001. Environmental Policies for Agricultural
Pollution Control (eds J.S. Shortle and D.G. Abler) 163