Chapter 2. The political economy of environmental and biodiversity relevant policy reform: Key obstacles and examples

Introduction

Political economy analysis attempts to examine the political factors that prevent markets, and governments, from achieving efficient outcomes (World Bank, 2008). While there is no single definition of what constitutes “political economy”, some of the more salient issues that arise in the political economy of environmental policy are competitiveness, income distribution, vested interests, as well as the political acceptability of reform.1 These issues are examined in turn below and examples of where these are relevant to biodiversity are reviewed.2

2.1. Competitiveness

A commonly cited obstacle to reforming environmental policies is the potentially adverse impacts on competitiveness. This can, in theory, manifest in two ways. First, intensifying environmental stringency will cause firms to incur higher compliance (production) costs, which drives up prices, reduces sales and profit, and can therefore result in at least some decrease in employment and economic health (Morgernstern et al., 2002).

Second, more stringent regulations may cause a competitive disadvantage compared to jurisdictions with lower standards, resulting in a loss of national competitiveness if nationally important sectors or firms are affected (Barker and Kohler, 1996). This creates an incentive for businesses to relocate to these low-standard jurisdictions (i.e., the so-called “pollution haven” effect) (Esty and Geradin, 1998; GFC, 2010), further affecting employment and national competitiveness. The pollution haven dynamic can influence policymaking, creating pressure to ease regulations in an effort to attract investment (the “race to the bottom” effect) or a “regulatory drag” whereby stringency is never increased (Esty and Geradin, 1998).

It is important to note that this does not necessarily mean that the economy as a whole will suffer, but rather that certain sectors may be negatively affected, while others are positively affected3 (OECD, 2006; Barker and Kohler, 1996).

The literature finds little empirical evidence of environmental regulation causing major economic or job losses (Morgenstern et al., 2002; Barker and Kohler, 1996; Shapiro and Irons, 2011; Albrizio et al., 2014a; Albrizio et al., 2014b) nor mass industry relocations (Konisky, 2007; Woods, 2006; GFC, 2010; Esty, 2011).4 Rather, relocations have been mostly driven by other factors such as lower labour costs, emerging markets, or lower corporate income taxes (Repetto, 1995;5 UNESCAP, 2012;6 Shapiro and Irons, 2011). Other literature suggests there may be small economic losses in the short term, as well as some industry relocation (Lenzen et al., 2012;7 OECD, 2006;8 Dechezleprêtre and Sato, 2014).

Woods (2006) suggests that the effects of regulation may be more pronounced for certain industries (e.g. chemical manufacturing) and for small start-up firms, although the specific impacts will depend on the industry’s sensitivity to price, the cost of abatement technologies, and the potential for jobs in abatement to replace any lost in the industry itself (Morgernstern, Pizer, and Shih, 2002).

Examples of studies examining competitiveness impacts in sectors or areas more closely related to biodiversity suggest mixed results. ECOTEC (2001) studied the influence of environmental taxes and charges in areas including water abstraction, wastewater, pesticides, fertilisers, and landfills, and found minimal effects on domestic and international competitiveness or on employment. Its case study of a pesticide tax in Sweden found that the tax comprised an average of 8% of the pesticide price, and had likely led to a decrease in pesticide consumption without causing any competitiveness impacts. In Denmark, the same was true even though the pesticide tax comprised a much higher average (37%) of the retail price. While possible adverse impacts on competitiveness have been frequently raised by farmers in the case of the French pesticide tax (see Chapter 3), the current pesticide tax is still very low, comprising about 5% of the pesticide price. In another study on the impact of the introduction of an effluent water tax in the Philippines, while there was some concern regarding potential relocation of firms, this did not appear to manifest in the end (Box 2.1).

Competitiveness issues may, however, be more relevant in other sectors related to biodiversity. High seas fisheries in particular are considered to be economically unviable without subsidies (GOC, 2013), suggesting that competitiveness losses can be expected by fleets of any nation that decreases subsidies while others do not. Reform can thus be hampered by a lack of international co-ordination, as nations wish to avoid putting their fleets at a disadvantage. Nevertheless, there are examples of successful fisheries reform in the cases of New Zealand (Box 2.2) and Iceland. In the case of Iceland, in addition to the introduction of a successful ITQ system, the government more recently also supplemented this with a resource rent tax to help address distributional concerns related to the initial free allocation of quotas (see Chapter 6).

In terms of empirical evidence for the “race to the bottom” phenomenon, Richards (2003) finds evidence for its existence in tropical forestry regulations, where it is catalysed by trade liberalisation, especially in countries with weaker forest governance structures. Woods (2006) also finds some indication of its occurrence in several U.S. states in the context of surface mining regulations, even in the absence of evidence for major investment movement: regulators simply acted on the potential that capital may move.

Appropriate compensating policies can help to counter negative effects on sectoral competitiveness. For example, recycling revenues back to business can offset increased compliance charges (Ekins and Speck, 1999; OECD 2006), as can applying border tax adjustments or countervailing duties to level the playing field for imports and exports (Esty and Geradin, 1998; OECD, 2006). Further, even where a sector’s output decreases, if compensation is provided by e.g. lowering labour costs such as employers’ social security contributions, overall employment levels may not change (Johnstone and Alavalapati, 1998). Sterner and Hoglund (2006) examine revenue recycling using the tax on NOx emissions in Sweden. They find that lower net tax payments as a result of recycling reduces resistance from polluters and make it politically easier to implement a tax rate high enough to provide abatement effects. Revenue recycling has also been used as a way to address concerns from the agricultural sector given the introduction of the French pesticides tax (see Chapter 3).

2.2. Distributional impacts

The expected distribution of costs and benefits of a policy is another important determinant of its political feasibility. Concerns surrounding regressive impacts of environmental policy reform have posed a barrier to progress (Johnstone and Alavalapati, 1998; Kerkhof et al., 2008; Wier et al., 2005; OECD, 2006). Income regressivity occurs when policy changes raise the price of basic goods, which forces low-income households to allocate a higher proportion of their budget to the goods than households higher on the economic scale.

The distributional effects of environmentally related taxes can arise from a variety of channels and have been broadly categorised as (OECD, 2006):

• The direct distributional effects on households arising from payment of the tax.

• The indirect distributional effects i.e. from price increases on taxed products from firms.

• The effects arising from the use of environmental tax revenues.

• The effects relating to benefits of environmental improvements.

As noted by OECD (2006), most studies on the income distributional impacts of environmental taxes tend to focus on energy and carbon taxation. Analyses of motor fuel and energy taxes (e.g. Bach et al., 2001; EEA, 2011; Fullerton and Heutel, 2007; Kerkhof et al., 2008; Tiezzi, 2005; Wier et al., 2005) find that, in general, the lowest income households bear larger increases in cost incidence.

In terms of policy instruments more directly relevant to biodiversity, the literature indicates mixed results. ECOTEC (2001), for example, found little evidence for concern regarding the distributional impacts of environmental taxes and charges. In contrast, in the Netherlands, two key issues contributing to the cancellation of a groundwater tax in 2011 were that the incidence fell predominantly on a narrow group of tax payers (10 drinking water companies paid nearly 90% of the tax), and that it did not effectively target environmental outcomes (Schuerhoff et al., 2013). In the UK, water charges caused regressivity due to pricing structures unrelated to the quantity of water consumed (EEA, 2011).

According to Von Moltke (2014), in most cases subsidies to fisheries are granted on the basis of valid policy goals and legitimate motives, including promoting development and poverty alleviation. Artisanal fishermen in Ghana, for example, are strongly against fuel subsidy reform because subsidy removal is expected to further increase fishing costs and thereby exacerbate poverty in their communities (Tanner et al., 2014). For payment for ecosystem services (PES) programmes, obstacles that have been encountered with regard to full empowerment of the poor may include high transaction, investment, and education costs; changing food or fuel prices; the opportunity costs of other livelihoods; the exclusion of informal land tenure; or elite capture (OECD, 2013). These are usually caused by inadequate attention to safeguards for empowering smallholders, agroforestry tenants, and ensuring an equitable distribution of benefits.

Concerns about distributional implications have also been raised in the context of individually transferrable quotas (ITQs). Such concerns in the U.S. for example, led to a moratorium on moving additional fisheries into ITQ programmes that lasted from 1996 to 2004 (Chu, 2008).9 However, a study by Brandt (2005) examining the equity implications of a regulatory change from command-and-control approaches to ITQs for the mid-Atlantic clam found that no segment of the industry was disproportionately adversely affected by the change. Similarly, in New Zealand, a study notes: “While it is clear that the number of small fishers has fallen since the introduction of the QMS (New Zealand’s ITQ-based “Quota Management System”) it appears that they, as a group, have been successful in finding alternative employment. From an employment perspective there is no evidence that New Zealand fishers have experienced significant social costs of restructuring the fishery.” (Stewart et al., 2006).

Finally, it is also important to ensure that any removal of subsidies does not lead to unintended outcomes that may be worse. For example, small-scale farmers are often dependent on agricultural subsidies, and may thus be forced to expand slash-and-burn agriculture when subsidies are removed, thereby causing more deforestation (Shandra et al., 2011).

In cases where the distributional impacts are likely to be a concern, appropriate policy packages can help to ease the transition. Recycling the revenue raised e.g. from taxes or subsidy removal through income or labour tax reductions can reduce, and may almost completely eliminate, the distributional effects, depending on the chosen method of implementation (Johnstone and Alavalapati, 1998). For example, the regressive impacts of water charges in Spain are offset by the progressive impacts of motor fuel taxes (EEA, 2011).

In Indonesia, the successful removal of pesticides subsidies, despite strong political opposition, has been attributed to taking advantage of fiscal or other policy crises (“policy windows”) to improve reform outcomes, and to supporting programmes and political economy conditions (e.g. decentralisation or support for budget reductions) that aided in implementation and in gaining popular support (CBD, 2011) (Box 2.3).

2.3. Vested interests and rent seeking behaviour

The influence of vested interests and rent seeking behaviour has also been cited as hindrances to environmental fiscal reform (Robin et al., 2003). Politically, reform “often involves trading off the concentrated benefits of vested interests against greater, but more widely dispersed, benefits to the public at large” (OECD, 2007) – in other words, calls for reform are likely to be far more dispersed than pressure to maintain the status quo. It has been noted for example that efforts to place a cap on biofuels in EU renewable energy targets, due to concerns of biofuel production causing increased deforestation, have been subject to heavy lobbying by the biofuels industry, causing legislative delays (EurActiv, 2013). In the case of fisheries, the overall lack of subsidy reform has also been attributed, at least in part, to interest group politics, since the fishing industry often enjoys a high level of influence domestically and in regional fishery management organisations (RFMOs) (GOC, 2013).

A number of other cases that have direct or indirect implications for biodiversity have been documented. For example, Swinnen (2010) finds that multiple EU Common Agricultural Policy (CAP) reform efforts have been impeded by farm interests, and American farming associations have been found to “significantly influence” agricultural subsidies through lobbying (Alvarez, 2005). American sugar producers have been so effective at mobilising lobbying networks that they have “[generated] unmatched levels of legislative support”, which has enabled them to enjoy increased protection even as other agricultural support mechanisms are dismantled (Alvarez, 2005). In Florida, this has resulted in the diversion of water from the Everglades; when returned to the watershed, it is loaded with fertiliser run-off that causes eutrophication (Robin et al., 2003). In Belgium, lobbying-driven exemptions for farmers from a pesticide tax led to the failure of the tax (ECOTEC, 2001). Also, in India, states in which wealthy farming groups hold major influence over the political agenda find it difficult to increase electricity prices to prevent groundwater overuse (World Bank, 2005).

Furthermore, “many reforms are designed to reduce or eliminate the rents accruing to small groups of privileged interests … however, these are precisely the policies that are most likely to be fought” (World Bank, 2008). The inflated rents reaped by these entities as a result of their efforts derive not only from capture of subsidies or grants, but also from lowered taxes, less stringent investment regulations (OECD, 2007), and foregone valuations of ecosystem services and biodiversity. The case study of Iceland (Chapter 6) is an example of a recent policy reform (via the additional introduction of a resource rent tax) specifically designed to allocate a greater share of the resource rents from the harvesting of a common property resource to the general public.

The resources at the disposal of rent seeking parties also allow them to be well focused, especially as compared to opponents who do not have the time or money to organise as efficiently (OECD, 2007). They may manipulate the legislative process to obtain advantageous results, which helps them create increased revenue and pressure the political process even more: “[subsidies] themselves create a pool of money out of which recipients can influence the very political process that channels money to them in the first place” (Steenblik, 1998). In Laos for example, illegal rent seeking is pervasive due to the structure of the permitting process and estimates suggest that bribes for permit approvals comprise logging companies’ single largest expense (Baird, 2009).

2.4. Political acceptability of reform

Increasing the stringency of environmental regulations or eliminating harmful subsidies is a process subject to complex political considerations that increase the difficulty of obtaining support. Societal conditions may influence the behaviour of elected officials, who feel the need to provide positive economic news; maintaining the status quo thus becomes politically attractive (OECD, 2005) and they may relax regulations or block more stringent policies to encourage business (Konisky, 2007). Other acceptability problems include distrust of government (OECD, 2006; Withana et al., 2012) and conceptual barriers: the public may not trust that the proposed tool will be as effective as claimed, or that the government will not appropriate revenue for other uses. In some cases, even though awareness may be high, the public may have trouble understanding the solution – e.g. a shift of taxes from income to pollutants (OECD, 2006).

Political acceptance is also dependent on (among other concerns) the perceived effectiveness of the policy, the degree of fairness, and the degree of awareness of the problem being addressed (OECD, 2006). For example, Soderholm and Christiernsson (2008) examine the political acceptance of fertiliser taxes in Austria, Denmark, the Netherlands, Norway, and Sweden. They find that the choice of tax scheme design matters not only for the cost effectiveness of the policy, but can also be an important mean of reducing any political opposition towards environmental taxes. They state: The European experience in fertiliser taxation indicates that some kind of earmarking of tax revenues can be effective in increasing the legitimacy of the tax policy, and taxes which achieve a close proportionality to damage done will often be perceived as fair.

Fairness most often concerns distributional or competitiveness impacts, but often the public may not be aware of any compensating mechanisms, or does not understand that policies are typically intended to assign responsibility to those who have contributed to the problem. Awareness is also subject to several secondary factors, including visibility of the issue (plastic bag pollution vs. biodiversity loss, for example); the ability to create a convincing narrative; and the ability to deploy messaging effectively. Narratives in fact play a major role: environmentally harmful industries have been particularly adept at creating “false perceptions and fear of change” about potential societal damage (Withana et al., 2012), framing reforms as competitiveness-reducing initiatives and thereby decreasing acceptability. Similarly, subsidy recipients often exaggerate the benefits to society of the support they receive (OECD, 2007). In some cases, social mores may influence the level of support granted, as for agriculture, which is often seen as an “old” and “traditional” industry that has provided much employment (van Beers and van den Bergh, 2001) and is therefore deserving of ongoing protection. This is also the case in Switzerland.

Regional laws can also hamper domestic ability for reform, as is occasionally the case in the EU. For example, the Euroviginette Directive does not allow Member States to account for impacts on biodiversity, landscapes, forestry, water, or other natural resources in determining road charges; the requirement for unanimity in the Council on tax-related measures may restrict reform progress; and feed-in tariffs for firewood-based cogenerated electricity in Estonia, implemented in support of EU directives, may cause overharvesting (DEFRA, 2012; Withana et al., 2012).