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

This chapter summarises the salient political economy issues that arise in environmental policy reform generally and provides examples of how they may create barriers to biodiversity related policy reforms. It draws on a literature review to identify common obstacles to reform. These include: potential competitiveness impacts, concerns about the distribution of costs and benefits, the influence of vested interests and rent seeking behaviour, as well as the political and social acceptability of reforms.



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).

Box 2.1. Effluent water tax in the Philippines

The Laguna de Bay region spans approximately 3 800 square kilometres and hosts some 13 million inhabitants and 10 000 enterprises (Catelo et al., 2007), including Manila and several smaller cities. As the second-largest lake in Southeast Asia and the largest in the Philippines, it is critically important for fishing and aquaculture, irrigation, power generation, and many other purposes. However, conflicts between water use and allocation, growth and development, and water quality have subjected it to numerous pressures over the last decades, including population expansion, urbanisation, industrialisation, overfishing, deforestation, and garbage dumping. These have caused eutrophication, biodiversity loss, health impacts, and flooding, among other issues (Deltares, n.d.).

In 1997, the Laguna Lake Development Authority instituted an Environmental User Fee System (EUFS) to encourage companies to decrease oxygen pollution as measured by biochemical oxygen demand (BOD). The EUFS consists of a fixed fee to cover administrative costs along with a variable fee based on volume of discharge and BOD load in the discharge. Despite some regulatory inefficiencies, it appears to have caused an 88% decrease in BOD loading from direct discharges of regulated companies (CBD, 2011) between 1997 and 1999 and a continued significant decrease by 2002 (Catelo et al., 2007).

The EUFS was initially applied to only large polluters in specific sectors, but eventually expanded to include smaller establishments and industries as well as farms and some residences. Industry was somewhat resistant and some firms appear to have contemplated relocating, but reconsidered once they were informed the EUFS would eventually extend country-wide (GWP, n.d.). Further expansion is ongoing, and the EUFS is expected to eventually to cover all possible sources of BOD loading as well as other types of pollution (CBD, 2011).

Source: Catelo, M. et al. (2007); CBD (2011); DELTARES (n.d.); GWP (n.d.).

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).

Box 2.2. Successful fisheries and agriculture reform in New Zealand


Subsidy reduction can be packaged with other fundamental policy changes or combined with other changes to the regulatory environment governing an industry to ease the adjustment process. In the case of fisheries in New Zealand, the financial crisis in the mid‐1980’s created favourable conditions for a major shift in policy towards the sector in the early 1990s. Subsidies were eliminated virtually overnight, a major change in management philosophy was introduced in the form of rights-based management and individual transferable quotas, and there was a minimum buy-out of existing rights. Subsidy reduction alone would not have been sufficient to create a sustainable fishing sector and would have caused substantial financial and social distress. It would also have an impact on fish stocks due to overfishing when fishers increase effort in order to try and cover marginal costs. In New Zealand, fishery subsidy reduction went hand in hand with a shift to a management regime (individual transferable quotas) which helped give those remaining in the fisheries sector a good chance at creating a profitable business environment, while allowing those who wished to leave to be bought out by those who wished to remain. As a result, fish stocks were managed more effectively and in some cases recovered from overexploitation.

Source: OECD (2007).


In the early 1980s, New Zealand began phasing out its agricultural subsidies, as part of an overall strategy to deregulate and otherwise reform public policies affecting key economic sectors (Myers and Kent, 2001). Price support, input subsidies, interest rate concessions, fertiliser subsidies, tax credits, and below-cost services provided by the Ministry of Agriculture were all removed (Johnson, 2001). While there were initial fears of the economic consequences of subsidy removal, farming productivity has actually been growing faster than before subsidy removal (Humphreys et al., 2003).

In the first years of the reforms, rural unemployment rose and farm incomes declined (OECD, 1998). However, land prices reverted to levels comparable to earnings (Johnson, 2001), and incomes eventually recovered: by 1990-92 the real net income for sheep and beef farms had returned to 1984 levels. Export earnings were maintained, though the composition changed (Johnson, 2001) and rural industries shed a significant percentage of jobs (OECD, 1998); nevertheless, the size of the agricultural sector has increased by 40 percent in constant dollar terms since 1986 (Humphreys et al., 2003). Fertiliser use decreased, as did capital expenditure on new plants and equipment; and a reduction in sheep numbers from 70 to 47.4 million led to land reclamation by woodland and shrubbery, improving ecosystem health. Recovery was also aided by diversification into areas such as value-added agricultural production, tourism, small business, technology, horticulture, fruit and vegetable production, and other local industries (OECD, 1998). Total factor productivity growth was approximately 2.5% annually from 1984-2007, compared to 1.5% annually before reforms (Gilmour and Gurung, 2007); agricultural productivity growth increased from 1 percent to 5.9 percent a year (Humphreys et al. 2003), and the effective rate of assistance decreased to -2% of agricultural production compared to a level of 38%, on average, between 1979-1983, indicating that the sector is now taxed (OECD, 1998). In 1984, the total area of private planted forest area was 500,000 hectares, while in 2001 it reached 1.7 million hectares (Humphreys et al., 2003), in part due to land converted from sheep and beef production to sustainable forestry.

Source: Gilmour, B. and R. Gurung (2007); Humphreys, J. et al. (2003); Johnson, R. (2001); Myers, N. and J. Kent (2001); OECD (1998).

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).

Box 2.3. The removal of pesticide subsidies in Indonesia

During the 1970s and 1980s, Indonesia’s agricultural policy promoted the use of high-yield crops and associated pesticides through direct subsidies on sales, credit concessions, and government spraying. Excessive usage of these pesticides in rice production caused biodiversity, health, and crop degradation, culminating in USD 1.5 billion worth of rice crop losses in the mid-1980s (CBD, 2011). These were a direct result of brown planthopper infestations, caused by pesticides eradicating the natural insect diversity that kept it in check.

In response, the government removed pesticide subsidies in 1986, and followed by ending fertiliser subsidies in 1996. Although there was strong initial resistance from some farmers, the government simultaneously introduced an integrated pest management system to train farmers in alternative pest control methods, and decentralised agricultural research to the provincial level (CBD, 2011). This helped to create a conducive policy environment for enacting the subsidy reforms, although it should be noted that budget stresses caused by oil price decreases also helped justify cuts (CBD, 2011). The policy package has been very successful: evidence indicates that the combination of technology dissemination and subsidy elimination reduced pesticide demand by 50 percent, saving the government some USD 100 million, while rice production still grew by three million tonnes over the next four years (de Moor and Calamai, 1997), though some of the growth can be attributed to increased fertiliser use (Gallagher, 2001).

Source: CBD (2011); De Moor, A. and P. Calamai (1997); Gallagher (2001).

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).


Acosta, A.M. and Jethro Petit (2013), “Practice Guide: A Combined Approach to Political Economy and Power Analysis”, Work in Progress Paper, Institute of Development Studies.

Albrizio, S. et al. (2014a), “Do environmental policies matter for productivity growth?: Insights from new cross-country measures of environmental policies”, OECD Economics Department Working Papers, No. 1176, OECD Publishing, Paris,

Albrizio, S. et al. (2014b), “Empirical evidence on the effects of environmental policy stringency on productivity growth”, OECD Economics Department Working Papers, No. 1179, OECD Publishing, Paris,

Alvarez, J. (2005), “Sweetening the US legislature: the remarkable success of the sugar lobby”, The Political Quarterly, Vol.76/1, pp. 92-99.

Bach, S. et al. (n.d.), The Effects of the Ecological Tax Reform in Germany, DIW Berlin, (accessed 28 November 2014).

Baird, I. (2009), Quotas, Powers, Patronage, and Illegal Rent-Seeking: The Political Economy of Logging and the Timber Trade In Southern Laos, Forest Trends.

Barker, T. and J. Köhler (1996), “Environmental policy and competitiveness”, Environmental Policy Research Briefs, Number 6, The European Union Research Network on Market-Based Instruments for Sustainable Development.

Brandt, S. (2005), “The equity debate: distributional impacts of individual transferable quotas”, Ocean & Coastal Management, Vol. 48, pp. 15-30.

Catelo, M. et al. (2007), Impact Evaluation of the Environmental User Fee System: A Stakeholder Perspective, Economy and Environment Program for South East Asia.

CBD (2014), Global Biodiversity Outlook 4, Convention on Biological Diversity, Montreal, QC.

CBD (2011), “Incentive measures for the conservation and sustainable use of biological diversity”, CBD Technical Series, No. 56, Convention on Biological Diversity, Montreal, QC.

Chu, C. (2008), "Thirty years later: The global growth of ITQs and their influence on stock status in marine fisheries", Fish and Fisheries, Vol. 10/2, pp. 217-230.

Colchester, M. et al. (2006), Promised Land: Palm Oil and Land Acquisition in Indonesia – Implications for Local Communities and Indigenous Peoples, Forest Peoples Programme, Moreton-in-Marsh, England.

De Moor, A. and P. Calamai (1997), Subsidizing Unsustainable Development: Undermining the Earth with Public Funds, Earth Council.

De Gorter, H. and J. Swinnen (2002), “The political economy of agricultural policy”, in Handbook of Agricultural Economics, Volume 2, B. Gardner and G. Rausser (eds.), Elsevier Science B.V.

Dechezleprêtre, A. and M. Sato (2014), The Impacts of Environmental Regulations on Competitiveness, Policy Brief, Grantham Research Institute on Climate Change and the Environment, London School of Economics, London, UK.

DELTARES (n.d.), Sustainable Development of the Laguna de Bay Environment.

DFID (2004), “The political economy of fishery fiscal reforms”, Fiscal Reform in Fisheries, Key Sheet #6, Department for International Development, London, UK.

EEA (2011), “Environmental tax reform in Europe: implications for income distribution”, EEA Technical Report, No.16/2011, European Environmental Agency, Copenhagen.

Ekins, P. and S. Speck (1999), “Competitiveness and exemptions from environmental taxes in Europe”, Environmental and Resource Economics, Vol. 13, pp.369-396.

Esty, D. (2011), “Economic integration and environmental protection”, in The Global Environment, 3rd Edition, R. Axelrod et al. (eds), CQ Press.

Esty, D. and D. Gerardin (1998), “Environmental protection and international competitiveness: A conceptual framework”, Faculty Scholarship Series, Paper 445, Yale Law School, New Haven, CT.

DEFRA (2012), Incentive Measures and Biodiversity – Rapid Review and Guidance Development, Department for Environment, Food, and Rural Affairs, London, UK.

ECOTEC (2001), Study on the Economic and Environmental Implications of the Use of Environmental Taxes and Charges in the European Union and its Member States, Birmingham, UK.

EurActiv (2013), “Biofuels industry sent ’three mails an hour’ in ILUC lobby offensive”,,, accessed December 23, 2014.

Felder, S. and R. Schleiniger (2002), “Environmental tax reform: Efficiency and political feasibility”, Ecological Economics Vol.42, Elsevier, Amsterdam, pp.107-116.

Fullerton, D., and G. Heutel (2007), “The general equilibrium incidence of environmental taxes”, Journal of Public Economics, Vol. 91, pp. 571-591.

Gallagher (2001), Stopping Subsidies for Pesticides in Indonesian Rice Production, Sustainable Development International.

Gilmour, B. and R. Gurung (2007), “New Zealand agriculture policy review”, Policy Note, Agriculture and Agri-Food Canada.

GFC (2010), “Competitiveness and environmental tax reform”, GFC Briefing Paper, No. 7, Green Fiscal Commission, London, UK.

GOC (2013), “Elimination of harmful fisheries subsidies affecting the high seas”, Policy Options Paper #6, Global Ocean Commission.

Grainger, C. and D. Parker (2013), “The political economy of fishery reform”, Annual Review of Resource Economics, Vol. 5, pp.369-386.

GWP (n.d.), “Philippines: establishing an industrial wastewater effluent fee program, Laguna de Bay”, Case Study, No. 82, Global Water Partnership.)

Haggard, S. and S. Webb (1993), “What do we know about the political economy of economic policy reform?”, The World Bank Research Observer, Vol. 8/2, Washington, DC, pp.143-168.

Høj, J. et al. (2006), “The political economy of structural reform: empirical evidence from OECD countries”, OECD Economics Department Working Papers, No. 501, OECD Publishing, Paris,

Humphreys, al. (2003), “Greening farm subsidies”, RIRDC Publication No. 03/040, Rural Industries Research and Development Corporation, Canberra.

IMF (2013), Energy Subsidy Reform: Lessons and Implications, International Monetary Fund.

Indrarto, G. et al. (2012), “The context of REDD+ in Indonesia: drivers, agents, and institutions”, Working Paper 92, CIFOR, Bogor, Indonesia.

Johnson, R. (2001), New Zealand’s Agriculture Reforms and their International Implications, Institute of Economic Affairs, London.

Johnstone, N. and J. Alavalapati (1998), “The distributional effects of environmental tax reform”, Discussion Paper, No. 98-01, International Institute for Environment and Development, London, UK.

Kerkhof, A. et al. (2008), “Taxation of multiple greenhouse gases and the effects on income distribution”, Ecological Economics, Vol. 67, pp. 318-326.

Konisky, D. (2007), “Regulatory competition and environmental enforcement: Is there a race to the bottom?”, American Journal of Political Science, Vol. 51/4, pp. 853-872.

Kostka, G. (2014), “Barriers to the implementation of environmental policies at the local level in China”, Policy Research Working Papers, No. 7016, World Bank Development Research Group.

Leal, D. (ed.) (2010), The Political Economy of Natural Resource Use: Lessons for Fisheries Reform, The World Bank, Washington, DC.

Lenzen, M. et al. (2012), “International trade drives biodiversity threats in developing nations”, Nature, Vol. 486/7401, pp. 109-112.

Morgenstern, R. et al. (2002), “Jobs versus the environment: An industry-level perspective”, Journal of Environmental Economics and Management, Vol.43, pp. 412-436.

Myers, N. and J. Kent (2001), Perverse Subsidies: How Tax Dollars Can Undercut the Environment and the Economy, Island Press, Washington, DC.

NRC (1999), Sharing the Fish: Toward a National Policy on Individual Fishing Quotas, Committee to review individual fishing quotas, National Academy Press, Washington, DC.

OECD (2013), Scaling-up Finance Mechanisms for Biodiversity, OECD Publishing, Paris,

OECD (2008), Report on Implementation of the 2004 Council Recommendation on the Use of Economic Instruments in Promoting the Conservation and Sustainable Use of Biodiversity [ENV/EPOC/GSP/BIO(2008)1/FINAL], OECD Publishing, Paris.

OECD (2007), Subsidy Reform and Sustainable Development: Political Economy Aspects, OECD Publishing, Paris,

OECD (2006), The Political Economy of Environmentally Related Taxes, OECD Publishing, Paris,

OECD (2005), Environmentally Harmful Subsidies: Challenges for Reform, OECD Publishing, Paris,

OECD (1998), Agricultural Policy Reform and the Rural Economy in OECD Countries, OECD Publishing, Paris,

Repetto, R. (1995), Jobs, Competitiveness, and Environmental Regulation: What are the Real Issues?, World Resources Institute, Washington, DC.

Richards, M. et al. (2003), “Higher international standards or rent-seeking race to the bottom? The impacts of forest product trade liberalisation on forest governance”, Impact Assessment of Forest Products Trade in Promotion of Sustainable Forest Management, Background paper, GCP/INT/775/JPN, FAO, Rome.

Robin, S. et al. (2003), “Perverse Subsidies and the implications for biodiversity: a review of recent findings and the status of policy reforms”, Sustainable Finance Stream, 5th World Parks Congress, Durban, South Africa.

Schuerhoff, M. et al. (2013), “The life and death of the Dutch groundwater tax”, Water Policy, Vol. 15, pp. 1064-1077.

Shandra, J. et al. (2011), “World Bank lending and deforestation: a cross-national analysis”, International Sociology, Vol. 26/3, pp. 292-314.

Shapiro, I. and J. Irons (2011), “Regulation, employment, and the economy”, EPI Briefing Paper, No. 305, Economic Policy Institute, Washington, DC.

Swinnen, J. (2010), “The political economy of the most radical reform of the Common Agricultural Policy”, Supplement: Perspectives on International Agricultural Policy, German Journal of Agricultural Economics, Vol. 59.

Steenblik, R. (1998), "Previous multilateral efforts to discipline subsidies to natural resource based industries", Report of Proceedings on Workshop on the Impact of Government Financial Transfers on Fisheries Management, Resource Sustainability, and International Trade, Vol. 17.

Sterner, T. and L. Isaksson (2006), “Refunded emission payments theory, distribution of costs, and Swedish experience of NOx abatement”, Ecological Economics, Vol. 57, pp. 93-106.

Soderholm, P. and A. Christiernsson (2008), “Policy effectiveness and acceptance in the taxation of environmentally damaging chemical compounds”, Environmental Science and Policy, Vol. 11, pp. 240-252.

Sutinen, J.G. (2008), “Major challenges for fishery policy reform: a political economy perspective”, OECD Food, Agriculture and Fisheries Papers, No. 8, OECD Publishing, Paris,

Tanner, T. et al. (2014), “Political economy of climate compatible development: artisanal fisheries and climate change in Ghana”, IDS Working Paper, No. 446, Institute of Development Studies, London, UK.

Tiezzi, S. (2005), “The welfare effects and the distributive impact of carbon taxation on Italian households”, Energy Policy, Vol. 33, pp. 1567-1612.

UNESCAP (2012), “Fact sheet – environmental tax reform and competitiveness”, Low Carbon Green Growth Roadmap for Asia and the Pacific.

Van Beers, C. and J. van den Bergh (2001), “Perseverance of perverse subsidies and their impact on trade and environment”, Ecological Economics, Vol.36, pp. 475-486.

Von Moltke, A. (ed.) (2014). Fisheries Subsidies, Sustainable Development, and the World Trade Organisation, Routledge.

Wier, M. et al. (2005), “Are CO2 taxes regressive? Evidence from the Danish experience”, Ecological Economics, Vol. 52, pp. 239-251.

Withana, S. et al. (2012), Study Supporting the Phasing Out of Environmentally Harmful Subsidies, Institute for European Environmental Policy (IEEP), Institute for Environmental Studies – Vrije Universiteit (IVM), Ecologic Institute and Vision on Technology (VITO) for the European Commission – DG Environment, Brussels, Belgium.

Woods, N. (2006), “Interstate competition and environmental regulation: A test of the race-to-the-bottom thesis”, Social Science Quarterly, Vol. 87/1, pp. 174-189.

World Bank (2008), The Political Economy of Policy Reform: Issues and Implications for Policy Dialogue and Development, Report No. 44288-GLB, World Bank, Washington, DC.

World Bank (2005), Environmental Fiscal Reform: WhatSshould be Done and How to Achieve it, World Bank, Washington, DC.


← 1. The literature on political economy of environmental reform covers several elements and sometimes uses different terms to describe similar issues. OECD (2006) for example examines competitiveness, income distribution, administrative costs, and political acceptance as the major political economy factors surrounding environmental taxes. De Gorter and Swinnen (2002) consider the influences of individual and politicians’ preferences, collective lobbying, and institutions, whereas special interest effects, voter ignorance, issue bundling, politician short-sightedness, de-coupling of costs and benefits, and bureaucratic inefficiencies are examined by Sutinen (2008). Other issues raised include administrative agency and discipline (World Bank, 2008; Haggard and Webb, 1993); crisis points (or lack thereof) (Haggard and Webb, 1993); external influences (Haggard and Webb, 1993); government trustworthiness or transparency (World Bank, 2008; DFID, 2004); income inequality (Høj et al., 2006); information deficiencies (IMF, 2013); institutional structures and rules (Acosta and Petit, 2013); property rights (Leal, 2010); reform timing (Haggard and Webb, 1993); strength and type of government (Haggard and Webb, 1993); trade (Leal, 2010); and vested interests and rent-seeking behaviour (World Bank, 2008; IMF, 2013; Haggard and Webb, 1993). Income distribution and income inequality are similar issues, as are crisis points and reform timing. This chapter addresses the most prominent issues addressed in the political economy of environmental reform.

← 2. It is important to note that comprehensive and comparable information on recent biodiversity-relevant policy reforms do not readily exist. Various studies have been undertaken to review progress in this domain such as the 2008 OECD Report on the Implementation of the 2004 Council Recommendationon the Use of Economic Instruments in Promoting the Conservation and Sustainable Use of Biodiversity. This found that although progress had been made with regard to subsidies that aim to promote biodiversity, much less progress had been made with regard to the introduction of instruments such as taxes, fees, and charges (i.e. those instruments that aim to internalise the negative external costs of production and consumption) or the reform of environmentally harmful subsidies (OECD, 2008a). A recent biodiversity tagging exercise of the OECD database on Policy Instruments on the Environment (PINE), finds that there are currently more than 400 instruments in place that are biodiversity relevant in the OECD and 35 non-OECD countries that provide information to this database.

← 3. For example, increased employment may be created in pollution-abating industries or in firms which are more easily able to comply with the regulations (Shapiro and Irons, 2011; Morgenstern, Pizer, and Shih, 2002; Dechezleprêtre and Sato, 2014).

← 4. These studies tend to focus on pulp and paper mills, plastic manufacturers, iron and steel mills, food and food products, chemicals, cement, and other manufacturing industries.

← 5. For the manufacturing sector, particularly pulp and paper, petroleum; organic and inorganic chemicals; coal mining; fertilisers; cement production; ferrous and non-ferrous metal production; metal manufacturers; and wood producers. Relocations due to labour cost differences occurred in industries such as textiles, apparel, footwear, and other light manufacturers, or producers of consumption items such as Coke.

← 6. For EU carbon and energy taxes.

← 7. For commodities directly impactingbiodiversity, such as coffee growing or forestry. Also see Meyfroidt, P. et al. (2013), “Globalization of land use: distant drivers of land change and geographic displacement of land use”, Current Opinion in Environmental Sustainability, Vol. 5/5, pp. 438-444.

← 8. For steel or other carbon-intensive production.

← 9. A report from the National Research Council was requested, and subsequently delivered in 1999, which recommended that Congress should lift the moratorium (NRC, 1999).