CHAPTER 2

LITERATURE REVIEW ON RIVER WATER QUALITY

PARAMETERS AND POLLUTION MANAGEMENT POLICIES

2.1 Purpose

This chapter aims to determine the current state of knowledge on river pollution

management by discussing relevant water quality parameters and their significance in

relation to river pollution assessment. It also reviews available policy instruments for

river pollution management to better understand their implications.

2.2 Water quality parameters

The health of a river depends on the quality of its water, which is influenced by the

presence of pollutants. The quality of water is generally assessed by a range of

parameters, which express physical, chemical and biological composition of water

(Meybeck and Helmer 1992). This research deals with some specific water quality

parameters of the Buriganga River, which include: temperature, pH, dissolved oxygen

(DO), biochemical oxygen demand (BOD), chemical oxygen demand (COD), electrical

conductivity (ECw) and the concentration of lead, chromium, ammonia nitrogen and

phosphate phosphorus. The guidelines as provided in the literature (Chapman and

Kimstach 1992; Liston and Maher 1997) have facilitated the appropriate selection of

these water quality parameters, which have been considered in relation to water use and

pollution sources.

2.3 Significance of selected water quality parameters

2.3.1 Temperature

Water temperature plays a significant role in affecting physical, chemical and biological

processes in water bodies (including the flowing waters like rivers), and thus the

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concentration of many variables (ANZECC 1992). High water temperature activates the

rate of chemical reactions with effect to evaporation and volatilisation of substances from

water. As water temperature increases, the solubility rate of gases in water such as

Oxygen (O2) decreases. Moreover, the respiration rates of aquatic organisms increase in

warm water which lead to greater consumption of O2 and increase the rate of

decomposition (Chapman and Kimstach 1992). An abrupt change in water temperature

can lead to greater destruction of aquatic life. On the other hand, excessively high water

temperature may lead to the problem of unwanted growth of water plants and wastewater

fungus (Metcalf and Eddy 1991). Surface water temperature can be influenced by factors

such as geographical position, seasonality, diurnal period, circulation of air, quantity of

cloud cover, depth of water and its flow rate. In general, the temperature of surface water

varies within the range of 0 0C to 30 0C; however, abnormally high temperatures can arise

from discharges of industrial effluent and sewage treatment plants (Chapman and

Kimstach 1992).

2.3.2 pH

The pH is used to read the acid balance of a solution and it is defined as ‘the negative of

the logarithm to the base 10 of the hydrogen ion concentration’ (Chapman and Kimstach

1992, p.62). The pH scale ranges from 0 to 14 (i.e., very acidic to very alkaline), and pH

7 indicates a neutral condition. The pH of natural water stays in between 6.0 and 8.5 but

could be affected by chemicals entering the waterways (Chapman and Kimstach 1992).

This is a significant parameter to assess water quality as it has influence over ‘many

biological and chemical processes within a water body and all processes associated with

water supply and treatment’ (Chapman and Kimstach 1992, p.62). This parameter can be

used to evaluate the amount of effluent plume in the water body, while measuring the

effects of an effluent discharge (Chapman and Kimstach 1992).

Extremely high or low pH values of fresh water make it unsuitable for most aquatic

organisms. Moreover, water with low pH values become corrosive to both metallic and

concrete structures in the water course and also reduce the availability of nutrients such as

calcium and magnesium. On the other hand, water with high pH values reduces the

availability of phosphate, sulphate, iron and manganese (Gambrell and Patrick 1988;

Jackson et al. 1993; Handreck and Black 1994). Furthermore, at high pH levels most of

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the dissolved carbon dioxide is converted into bicarbonate (HCO3⎯) or carbonate (CO32⎯)

(Sand-Jensen and Gordon 1984; Larkum et al. 1989).

This parameter has a direct effect on the treatability of the wastewater by biological

means, particularly in the control of aerobic digestion of organic matter. For example, if

the pH approaches to 5.0, the acidic stage of digestion is becoming predominant and

digestion will be unsatisfactory. A pH value varying between 6.5 to 8.0 is required for a

proper biological treatment of wastewater (Metcalf and Eddy 1991).

2.3.3 Dissolved oxygen

The analysis of dissolved oxygen (DO) is used to measure the amount of gaseous oxygen

dissolved in the water, which is crucial for all forms of aquatic life. DO in water mainly

appear by diffusion from the atmosphere and also from the photosynthesis of aquatic

plants. Determination of this parameter is an integral measure of assessing water quality

as O2 plays an influential role in ‘nearly all chemical and biological processes within

water bodies’ (Chapman and Kimstach 1992, p.65).

The DO measurements can help to determine the level of pollution caused by O2

demanding substances, such as biodegradable organic matter and nutrients (Masters

2004). The effect of O2 depleting substances along the length of the river can be

illustrated with the ‘oxygen-sag curve’ (Masters 2004, p. 131), and this depletion can be

observed from a few kilometres to 100 km downstream of the discharge point (Figure

2.1). The eventual recovery in O2 concentrations is enhanced by high water turbulence in

the flowing water body (Chapman and Kimstach 1992; Masters 2004). The determination

of DO is also used in the measurement of biochemical oxygen demand (BOD) and

chemical oxygen demand (COD) in water (discussed in following sections).

The concentration of DO in an aqueous system can change, subject to the season or even

within a 24 hour time period, in terms of temperature and biological activity

(photosynthesis and respiration). In general, as the temperature and salinity level in the

water increase, consequently, the level of DO decrease (Goldman and Horne 1983). In

fresh water, DO at sea level ranges from 15 mg/L at 0 0C to 8 mg/L at 25 0C. DO can also

be measured in terms of percentage saturation, and the saturation levels below 80 percent

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in drinking water are identifiable in terms of taste and smell (Chapman and Kimstach

1992). Stagnant water usually contains less DO levels than flowing water because of

turbulence, which creates more opportunities for circulation of O2 across the air-water

interface (Xu 2006).

Adapted from: Arceivala 1981

Figure 2.1. Typical changes in DO downstream of a wastewater input to a river

2.3.4 Biochemical oxygen demand

The biochemical oxygen demand (BOD) is used to read the level of biochemically

degradable organic matter or carbon loading in the water. It is usually defined by the

amount of O2 consumed by the aerobic micro-organisms present in the water sample for

the purpose of oxidising the organic matter and to convert it to a stable inorganic form

(Chapman and Kimstach 1992; Liston and Maher 1997). Hence, in water quality analysis

this parameter is used to determine the biodegradable organic content of the waste in

terms of O2 which is required when the wastes are discharged into natural water where

aerobic condition prevails.

As the wastes which contain biodegradable matter are released into a body of water,

microorganisms, especially bacteria, feed on the wastes and break them down into simple

organic and inorganic substances. During this decomposition in an aerobic environment,

the process produces stable end products such as carbon dioxide (CO2), sulphate (SO4),

orthophosphate (PO4) and nitrate (NO3). The process can be represented by the following

form (Masters 2004, p.117):

Organic matter + O2 CO2 + H2O + New cells + Stable products micro-organisms

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The BOD is usually determined through standardised laboratory procedures where the

sample is incubated in the dark at a steady temperature of 20 0C for the duration of 5

days, thereby measuring the amount of O2 consumed in this process. This explains the

term BOD5 (biochemical oxygen demand on five days). Unpolluted waters typically

contain BOD5 values of 2 mg/L or less, while raw sewage could have a BOD5 value of

about 600 mg/L (Chapman and Kimstach 1992). BOD5 values are often used as a robust

surrogate of the degree of organic pollution in water body as it can accurately depict real

world impact on receiving waters.

2.3.5 Chemical oxygen demand

The chemical oxygen demand (COD) is commonly used to measure the susceptible levels

of oxidation of the organic and inorganic materials existent in water bodies as well as in

the sewage and industrial effluents. It measures the O2 equivalent of the organic matter

present in a water sample that can be oxidised by a strong chemical oxidant, such as

dichromate or permanganate (Chapman and Kimstach 1992). The 2 to 4 hour laboratory

tests for COD measures the level of O2, which is necessary for chemical oxidation of

organic and inorganic matter in the water sample to convert into CO2 and water. The

COD test does not aim to identify the oxidisable material or find differences between the

organic and inorganic material in the water. However, it has been a widely used measure

for water quality analysis over the past several decades. The concentration of COD

observed in unpolluted surface water remain around 20 mg/L or less, while values are

normally greater than 200 mg/L in effluents (Chapman and Kimstach 1992). It is noted

that COD measurements are usually higher than the BOD5 measurements (Masters 2004).

2.3.6 Electrical conductivity

Electrical conductivity in water (ECw) is a measure of salinity and the extent to which

water is able to conduct an electric current. It is expressed as micro Siemens per

centimetre (µS/cm) and, relates to the concentrations of total dissolved solids (TDS) or

salts in a specific water body (Chapman and Kimstach 1992; Taylor 1993; Liston and

Maher 1997). These salts typically include such cations as sodium, calcium, magnesium

and potassium, and anions such as chloride, sulphate and bicarbonate (Masters 2004). The

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salinity or the TDS in mg/L of water may be calculated by multiplying the conductance

(in µS/cm) by a factor which is commonly used as 0.68 (Al Bakri and Chowdhury 2002).

‘The conductivity of most freshwaters ranges from 10 to 1,000 µS/cm’ (Chapman and

Kimstach 1992, p.62). However, if the value exceeds 1,000 µS/cm it renders adverse

biological effects in freshwater rivers, streams and wetlands (Chapman and Kimstach

1992). Moreover, higher values of ECw significantly affect the use of water for irrigation

and drinking purpose (Hart et al. 1990; Chapman and Kimstach 1992). In water bodies,

often ECw is influenced by the geology (rock composition) of the watershed. However,

there are a number of sources of pollutants, which may be signalled by increased ECw.

These include wastewaters from sewage treatment plants, urban runoff and industrial

plants (Masters 2004).

2.3.7 Heavy metals: lead and chromium

The presence of heavy metals in fresh water (more than the acceptable levels) is

responsible for serious ecological problems because of their toxicity. As the heavy metals

cannot be removed from water bodies through natural elimination process, they move

from one section of the aquatic environment to another, including the biota through food

chain transfers (Chapman and Kimstach 1992). This phenomenon of heavy metals often

cause detrimental effects within the ecosystem and increase toxicological risk for human.

Heavy metals have a range of adverse effects on the human body, including nervous

system and kidney damage, creation of mutations and inductions of tumours (Masters

2004). Thereby the measurement of heavy metal pollution has become a significant part

of most water quality assessment programs.

The list of heavy metals includes many chemical substances but this particular research

includes only tow metals namely lead (Pb) and chromium (Cr) for assessment purpose.

These two heavy metals have been primarily selected on the basis of the guideline

provided by Chapman and Kimstach (1992). Moreover, Pb and Cr are also included

within the ten priority listed metals of the Global Environment Monitoring System

(GEMS) program and eight high priority listed metals of the United States Environmental

Protection Agency (US EPA) (Chapman and Kimstach 1992).

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Lead (Pb) contamination in freshwater is caused by atmospheric deposition, industrial

discharge and leakage of oil. The other heavy metal, chromium (Cr), enters into water

from industrial effluents like tanneries, textiles, ceramics and chrome plating industries.

There are different chemical forms of Cr but this research will focus on the measurement

of chromate, Cr (VI), as in the aquatic environment Cr exists primarily in this form

(Sharman et al. 2008).

2.3.8 Nutrients: ammonia nitrogen and phosphate phosphorus

Nitrogen and phosphorus are fundamental elements of nutrients and their availability is

essential for plant and animal growth (Gundersen and Bashkin 1994). However, nutrients

in water are often identified as pollutants and are detrimental to water quality especially

when heavy concentration of nutrients fosters favourable conditions to expedite the

growth of unwanted aquatic plants, particularly algae. These nutrients are among the

factors that cause eutrophication in aquatic ecosystems (Metcalf and Eddy 1991; Liston

and Maher 1997). As the algae eventually grow and die, their decomposition decreases O2

in the water, potentially leading to reduced DO levels. Moreover, algae and other

decaying organic matter add colour, turbidity, odours and objectionable tastes to water

that are difficult to remove and that may greatly reduce its acceptability as a domestic

water source (Masters 2004). The major sources of both nitrogen and phosphorus in water

include municipal wastewater discharges, sewage, urban and agricultural runoff, animal

feed lots and industrial wastes. Detergents and other laundry materials are the major

contributors of phosphorus in water (Welch 1980; Masters 2004).

Water containing ammonia nitrogen (NH3-N) is normally considered as recently polluted

while those containing other forms of nitrogen are polluted earlier. Unpolluted water

contain small amounts of NH3-N, usually less than 0.1 mg/L, while higher concentration

could be a sign of organic pollution in water (Chapman and Kimstach 1992; Makepeace

et al. 1995). Phosphate phosphorus (PO43⎯-P) is the common dissolved form of

phosphorus, and it accurately indicates the level of phosphorus immediately available for

consumption by aquatic plants (Liston and Maher 1997). In most natural surface waters,

PO43⎯-P ranges from 0.005 to 0.02 mg/L. High concentration of PO4

3⎯-P can indicate the

presence of pollution and is the main cause of eutrophication in water bodies (Chapman

and Kimstach 1992; Makepeace et al. 1995).

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2.4 Literature review on policy instruments for river pollution management

Policy instruments for river pollution management comprise of a set of arrangements,

generally introduced by government, for pollution abatement to attain some target level of

water quality in a river (Perman et al. 1999). There are various kinds of direct and indirect

regulatory instruments that can be used to address the river pollution problems. A

considerable literature exists on the use of different policy instruments for pollution

control. Each of the instruments has their own advantages and disadvantages in various

stages of their application. Hence the review of the theoretical and practical aspects of

commonly used policy instruments assist to determine the most appropriate instruments

for possible implementation in the Buriganga River.

The approaches to mitigation of pollution can be grouped within two broad systems, such

as: (a) Government regulations; and (b) Non-government or self regulations. There are

also further classifications of instruments under each regulatory system. The government

regulations for pollution control can be categorised as (1) Command-and-Control based

instruments; and (2) Economic incentive based instruments. The following sections will

review the literature on selected pollution control instruments in terms of their nature,

efficiency, advantages and disadvantages of their use in river pollution control.

2.4.1 Command-and-Control based instruments

Command-and-Control (CAC) based instruments operate by enforcing direct regulations

on processes or products, by imposing acceptable levels on the emission of particular

pollutants, by issuing restrictions on polluting activities, and by limiting the polluters to

operate at particular areas and time (Chave 1997; Bernstein 1997). Hence, application of

CAC based instruments (direct regulations) heavily relies on setting up of various

quantitative and qualitative controls and regulations along with monitoring and

enforcement systems to limit polluters’ behaviour (Bernstein 1993; Kolstad 2000).

Generally, these direct regulatory measures consist of laws mandating emitters of

pollutants to ensure meeting a pre-determined uniform reduction level (Austin 1999). In

the CAC based approach, it is a prerequisite for the government to outline ambient

environmental objectives pertaining to health or ecological issues and to set specific

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standards on allowable levels of pollutant discharge. This approach also specifies the

technology the polluters should adopt to fulfil the environmental targets (Bernstein 1993).

Bernstein (1993) identifies that the direct regulatory measures empower the regulator

with the supreme authority to manipulate the utilisation of resources in order to meet a

given environmental objective. CAC based instruments result with a more certain level of

pollution abatement, and may be able to effect rapid improvements in ambient

environmental quality (Murty 1999; Kolstad 2000). Therefore, one of the major

advantages of this instrument is its reliability in bringing about a desired environmental

outcome within a short period. However, under a pure CAC regime, polluters may not

have adequate incentive and flexibility to invest in more effective pollution control

mechanisms or adopt cleaner process technologies on a long term basis (FCE 1994;

Murty 1999; Kolstad 2000).

Another major disadvantage of these instruments is their economic inefficiency (Baumol

and Oates 1988; Larsen and Ipsen 1997; Tietenberg 2006). For example, under the CAC

approach, the same emission standards apply to all commercial enterprises, irrespective of

their pollution abatement costs (Bernstein 1993). In this same vein Kolstad (2000)

verifies that the marginal costs of pollution abatement among different polluters

generating the same pollutant can not be equalised as an outcome of applying CAC based

instruments. Equalisation of marginal costs of pollution abatement could only occur if

regulators were completely correct in their assessment of each polluter’s abatement costs,

which is highly unlikely in practice (Perman et al. 1999).

CAC based instruments limit emissions from each source and allows no provision for the

sources to trade in their pollution rights (Eskeland and Jimenez 1992). Consequently,

some polluters are automatically forced to reduce the emissions at relatively higher costs

compared to the other polluters who are capable to reduce emissions at relatively lower

costs. By the early 1990s, regulators in many OECD countries came to the conclusion that

traditional CAC based instruments were too expensive and often ineffective for them to

use (Austin 1999; World Bank 2000). Nevertheless, in practice these instruments are still

the most prevalent and predominant form of environmental regulation throughout most of

the developing world (Eskeland and Jimenez 1992; Perman et al. 1999; Murty 1999). The

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following sections will further analyse the advantages and disadvantages of the CAC

based instruments in its various specific forms such as standards, permits and licenses.

Standards

Standards are the most common form of direct regulations of water quality. There are

different types of standards including ambient standards, effluent or emission standards,

technology-based standards, performance standards, product standards and process

standards. The function of the different standards is to set a reference for evaluation, goal

setting for legislative action and for pollution control (Bernstein 1993; Field 1994).

Larsen and Ipsen (1997) emphasised that water quality standards should always be

adopted considering the local (achievable) economic and technological level and the

assimilative capacity of the receiving water. Moreover, they also suggested that accurate

determination of standards is important to avoid inappropriate (under or over) treatment

of pollutants. Otherwise, inappropriately set standards may lead to needlessly high

treatment costs or to excessive pollution.

Ambient water quality standards indicate the minimum conditions to be met for particular

parameters and at particular locations according to the intended use of the water body

(Larsen and Ipsen 1997; Bernstein 1993). This type of standard enables to optimise

treatment efforts and costs because the level of treatment may be adjusted to the actual

assimilation capacity of the receiving water. However, Bernstein (1993) identifies a

significant problem with this approach to be its implementation, as the knowledge of the

assimilative capacity necessitates studies of the hydraulic, dispersive, physico-chemical

and biological conditions existing in the water body.

Effluent or emission standards call for specific level of treatment of all wastewater,

irrespective of the state of the receiving water body. They provide direct and manageable

means for controlling pollution from the sources and require relatively simple

administrative competence to measure the pollution concentrations at the source (Larsen

and Ipsen 1997). However, Fano et al. (1986) contradict this argument and mention that

the application of this standard may require enormous administrative and enforcement

costs for developing countries. Another issue regarding the application of emission

standards as noted by Bernstein (1993) is that it is usually conducted by government

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inspectors through spot checks. The inspectors are empowered to impose penalties on

violators. On the other hand, the violators may delay in complying with standards,

thereby involving the government in long standing legal processes. This often jeopardise

the effectiveness of the application of effluent or emission standards. Besides, Helmer

(1987) postulates that application of these instruments does not consider the water quality

requirements of the receiving water bodies. They can provide overprotection in some

rivers and insufficient protection in others.

Technology-based standards specify particular technologies that pollution generators

must avail to comply with environmental laws. They provide no flexibility to polluters to

determine the choice of technology to use in order to meet such requirements. However,

in reality different polluters could have different marginal abatement cost (MAC)—the

cost of eliminating an additional unit of pollution—as they perform differently to reduce

their pollution due to source-specific constraints. Hence, it may not be cost effective for

the polluters if they are forced to take similar abatement measures (Andersen 2001). On

the other hand, performance standards indicate the load of dischargeable pollutants,

focusing on the allowable discharge concentration, and specifying the amount of

pollutants that should be eliminated before being discharged (Bernstein 1993; Jaeger

2005). The advantage of performance standards is that it allows polluters to use the least

expensive technique in order to meet water or other environmental quality requirements.

However, OECD (1987) contends that these instruments are best suited mostly for large

and complex enterprises that have the freedom to select the most effective techniques for

pollution control.

Permits and Licenses

Pollution control permits and licenses can be effective for regulating point source

discharges. These instruments are usually linked to an environmental quality standard and

may depend on particular conditions such as compliance with a code of practice, selecting

location that could minimise negative impacts on the environment and the economy,

constructing treatment plants or adopting pollution control equipments within a specified

time frame, and implementing other protective measures to safeguard the environment

(Chave 1997). An advantage of permits and licenses is that they include all pollution

control obligations of an enterprise in one document. Moreover, permits and licenses

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provide controlling power to the government as they are liable to be suspended or totally

withdrawn any time subject to the demands of the national economy or societal interests

(Bernstein 1993). However, the use of permits and licenses usually involves regular

monitoring and reporting on the environmental performance of the enterprise.

2.4.2 Economic incentive based instruments

Economic incentive (EI) based instruments are also known as market-based instruments.

These instruments are shaped by market forces and they aim to change polluter behaviour

in favour of environmental conservation (Bernstein 1997; Hanley 1997). In the system of

applying these instruments, polluters are not told how much they can pollute or what

technology they must use, but their choices have financial consequences and hence this

influences the choices they make. With these policies, emission constraints are not

specific to a given source; rather they provide equal monetary incentives to all polluters

by effectively increasing the marginal costs of production (Oates and Baumol 1975;

Eskeland and Jimenez 1992).

Though the knowledge of the EI based instruments have existed in the literature along

side the more conventional regulations (such as the standards), their application in

pollution control has not been quite widespread (Austin 1999). However, with the

growing understanding of the advantages of the EI based instruments and with the

increased urgency to balance the cost of environmental protection with its benefits, it is

anticipated that the use of these instruments would be more widespread in the near future.

Hence the empirical assessment of the possibility of applying EI based instruments is

gaining importance in environmental decision making (Blackman 2009).

Probably the greatest advantage of EI based instruments is that they can minimise

pollution at a reduced total cost while achieving the same level of environmental

protection (Pearce and Turner 1990; Austin 1999; Jaeger 2005). Moreover, Bernstein

(1993) argues that as EI based instruments can control pollution according to market

mechanisms, they can deregulate and reduce government involvement. In general, proper

implementation of these instruments has several advantages widely advocated in the

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literature (OECD 1989; Bernstein 1993; Panayotou 1994; Larsen and Ipsen 1997; Murty

1999; Perman et al. 1999; Kolstad 2000; UNEP 2005), such as:

• promoting cost-effectiveness to reach acceptable pollution levels;

• encouraging innovative technological solutions for the purpose of reducing the

cost of pollution control;

• providing government with a revenue source to support pollution control

programs;

• providing flexibility to adopt any technology for pollution control; and

• relieving the government from conducting extensive studies to determine specific

level of control for each enterprise and its products.

Despite these significant advantages, there are certain disadvantages to the EI based

instruments. Compared to the traditional CAC based regulatory approach, the EI based

instruments provide less certain and less predictable outcomes on environmental

protection as the polluters have the liberty to choose their own measures (Bernstein

1993). As a result, regulatory agencies often feel reluctant to use these instruments as they

think that the EI based instruments may not empower them with more control over the

polluters or predict more accurately about environmental outcomes.

Moreover, these instruments are not appropriate for all types of pollution control or to

deal with environmental emergencies, as the rates at which instruments are applied are

not easy to change at short notice. Bernstein (1993) and Murty (1999) suggested that

these instruments cannot be quickly adjusted to deal with abnormal amount of emissions

arising out of emergencies, and even if the changes are achieved, the polluters’ responses

are not prompt. Further, even though EI based instruments reduce the government’s role

to some extent, Bell and Russell (2002) argue that these instruments require good

institutions, as a precondition, to effectively implement and enforce them. Bower et al.

(1981) point out that a system with EI based instrument is administratively and legally

more difficult to implement, and may be abused by polluters. This critique further adds

that the challenges associated with monitoring compliance are greater for EI based

measures. Therefore, the cost of implementation of EI based instruments may be higher

compared to other policies.

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For practical purposes, there are different types of EI based measures such as pollution

taxes/fees, marketable permits, subsidies and deposit refund systems. However, among

the various instruments, pollution taxes and marketable permits are more common (which

are also of primary interests for the present research) and have received recent attention in

many countries including developed and developing countries (Blackman 2009a). There

are various forms of these two instruments; hence, the advantages and disadvantages of

these two specific types of EI based measures are further discussed in the following

sections.

Pollution taxes

Bernstein (1997) states that there are four main types of pollution taxes which are used in

practice: (1) effluent charges (determined on the quantity and/or quality of the discharged

pollutants); (2) user charges (fees charged for using collective treatment facilities or

natural resources); (3) product charges (fees levied on environmentally harmful products

when used as ‘an input to the production process, consumed or disposed of’); and (4)

administrative charges (fees payable to authorities for such purposes as chemical

registration and licensing).

These charges or taxes are known as the price payable for the use of environmental

resources (such as the rivers) or to control the ‘incremental units of pollution’ (Bernstein

1993, p.10). In the literature these instruments are also termed as price based

mechanisms. The taxes on the polluters are imposed based on the idea of reducing the

negative externality (or the social costs) forced on society. This concept is directly related

to the paradigm of taxing pollution causing agents which was originally proposed by

Pigou (1912, 1932), and is also known as the Pigouvian tax. In the case of negative

externalities, Pigou proposed a tax on emissions as a corrective measure to encourage

economic activity by setting the tax rate on a par with the marginal external cost of the

pollution (Murty 1999). Theoretically, a system of Pigouvian tax offers incentive for the

polluters to reduce pollution, while in a direct regulatory system (such as the uniform

reduction measure), a pollution emitter does not receive any such incentive to reduce

pollution below the acceptable level.

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Based on Pigou’s principle, Baumol (1972) postulates that if the external cost, such as the

pollution is not internalised, markets may fail to achieve the optimal outcome. In practice,

setting of an optimal pollution tax could be a difficult and complicated task, as it requires

information about the exact quantity and quality of the discharged wastewater from

individual sources (Bernstein 1993; Kraemer 1995). Moreover, a downward pressure on

tax rates may be affected through political factors, such as when polluters lobby to the

government, which may tend to reduce the efficiency of this instrument. However, Afsah

et al. (1996) have pointed that regulators will be able to apply pollution taxes to manage

pollution more cost effectively when they have more integrated information systems

coupled with strong public support. Anderson et al. (1989) contend that this instrument is

more ‘appropriate when the damage from incremental units of pollution can be estimated

reliably’ (Bernstein 1993, p.10). Another important issue with the pollution taxes is that

the rates have to be periodically revised upwards or downwards depending on whether the

actual level of pollution is below or above the optimal level (Baumol and Oates 1988).

This adjustment also depends on the availability of information and on political situation

(Kolstad 2000).

Marketable permits

The fundamental idea underpinning the marketable permits was stated by Coase (1960)

who suggested that the harmful effects on nature ‘could be regulated as effectively and

efficiently by a market as by the more conventional forms of regulation’ (Ellerman 2005;

p.123). A quantity based measure was then developed and applied specifically to

environmental problems by Dales (1968) who proposed the idea of creating tradable

property rights for natural resources. Later, Montgomery (1972) advocated the need to

define a separate property right in order to determine the damage caused by a specific

source at all the downstream sites. This would enable to achieve an economically viable

goal.

In order to apply tradable permits, the authority (such as the government regulator) has to

set a cap on a pollutant in terms of permissible level of pollution. The authority can then

distribute (either based on grandfathering or through a primary auction) the total

permissible pollution load among the pollution emitters by issuing a permit system. This

enables the pollution emitters to discharge a certain amount of effluent over a certain time

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period. In this system permits are allowed to be traded among the polluters and they have

the flexibility to trade permits both externally (between different polluters) or internally

(between different plants within the same enterprise) (Baumol and Oates 1988;

Tietenberg 1990; Eskeland and Jimenez 1992; Bernstein 1993; Bernstein 1997; Murty et

al. 1999; Beaumont and Tinch 2004; Ellerman 2005).

In this mechanism, the polluters with the higher cost of abatement are more inclined to

buy permits than to reduce pollution. Since polluters have different marginal costs of

abatement, a market would be created where low cost polluters would sell permits and

high cost polluters would purchase them. Therefore, the opportunity to trade permits

among the polluters would minimise the total cost of pollution mitigation. This could also

lead to an eventual decrease of overall emission of pollution as the polluters whose

abatement costs are low may be more inclined to further reduce emissions with cheaper

abatement options (Eskeland and Jimenez 1992; Beaumont and Tinch 2004).

Despite having the important advantage of cost-effectiveness, there are certain limitations

of marketable permits, especially if they are applied in a developing country where

administrative and institutional systems are weak (Kathuria 2006). Also, application of

this instrument requires strict enforcement of the pollution cap, which will require

constant monitoring of both quantity and quality of pollutants discharged (Murty et al.

1999; OECD 2001). Moreover, these instruments work effectively in well-functioning

markets, where a large number of enterprises exist who are willing to either buy or sell

permits (Bernstein 1997). Otherwise, larger enterprises may exert some influence on

permit prices and the market may be incapacitated (Baumol and Oates 1988; Perman et

al. 1999).

Further, Stavins (1995) has pointed that it is important to keep the transaction costs low

for effective permit trading in the market. The transaction costs may rise in the absence of

a well-developed administrative system (for bargaining and decision) and inadequate

monitoring and enforcement capacity (Stavins 1995; O’Connor 1998). However,

Dahlman (1979) interpreted these problems as costs incurred due to lack of information

on abatement technologies and their prices. Hence, Tietenberg (1998a) emphasised on the

reduction of transaction costs by providing public information on prices for effective

implementation of such instruments. Kathuria (2006) has contended that the possible

25

(theoretical) advantages of using marketable permits may overcome the costs involved in

rectifying the limitations levelled against it. There are different types of marketable

permits in use (Hahn and Hester 1989; Stavins 2001; Ellerman 2005). The main types are

discussed below:

Cap and trade: This is the most widely used form of tradable permit. It provides permits

for emitting pollution (either at the source or at the receptor) and allows the participants

to trade those amongst themselves depending on the aggregate cap as set in the beginning.

In this system the polluters need to possess adequate permits to cover for the emissions

generated by them. The total numbers of permits cannot exceed the cap, which limits the

total emissions to the aggregate level. Polluters who need to increase their emission levels

have to purchase permits from those who require fewer permits. This system ensures that

the buyer pays a fee for generating more pollution, while the seller is being rewarded for

minimising emissions.

Offsets: These are used by new or modified pollution sources. The environmental

regulators require the new sources to offset their emissions by reducing emissions

elsewhere. The new sources are allowed to buy credits from existing firms. This

instrument is becoming increasingly popular in number of applications, such as salinity,

water quality and biodiversity.

Bubbles: These are used for reduction of emission of a particular pollutant in a specific

location. The bubble allows an increase in emissions from one polluter if another source

of the same pollutant within the bubble (a specific geographical area) reduces emissions

by the same amount.

Netting: This enables to avoid the most stringent emission limits that would be applied to

the modification by decreasing emissions from another source within the same

plant/enterprise. This reduces the net emission increase to a level below that is considered

significant.

Banking: It provides a mechanism for pollution generators to save emission credits for

future use for bubble, offset or netting programs.

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2.5 Literature review on informal instruments for river pollution management

There are some self regulatory instruments such as public participation, public disclosure,

and voluntary agreements which can complement the formal government regulations like

CAC or EI based instruments for pollution control (Chave 1997; Blackman 1999; Pu

2003; Blackman 2009b). In the literature, these instruments have been termed as

‘informal’ as their use for pollution control on their own is generally not sufficient and

they are recommended to be used only to complement the formal instruments. Moreover,

for policy development purpose, the instruments remain informal as long as they do not

receive the official recognition for their use in pollution control. However, O’Connor

(1998) termed public disclosure of information and voluntary agreements as suasive

instruments. Discussions on the implications of different informal instruments are as

follows:

Public participation in pollution control

Numerous international declarations (such as Dublin Statement 1992) have already

documented public participation as essential element for integrated water resources

management (ICWE 1992; Mostert 2006). UNEP (2002) has emphasised that public

attitude towards environmental pollution is an important component of a sustainable

strategy. Moreover, Healy (2006) has stated that in areas where much data on pollution is

not available, public involvement and viewpoints can identify previously unconsidered,

yet constructive strategies. Chave (1997) points out that the involvement of the public is

essential to enable regulators to understand the impact of any proposed measure prior to

setting standards for water and effluent, and to ensure that any programs for improvement

are attainable within the financial and technical capabilities of the country concerned.

However, it is apparent that the degree of association of the public in a pollution control

program would very much depend on the socio-economic characteristics of the local

community and their willingness to get involved.

The primary reason for community engagement is usually to inform and educate so as to

promote public understanding, agreement and perhaps achieve consensus regarding an

environmental problem. Bhushan (2004) observed that in many cases where community

pressure is absent, regulators and polluters share a privileged partnership. The pressure

27

from a strong knowledge-based community can become a driving force to push polluters

towards environmental protection. Pu (2003) argued that governments can strengthen the

power of communities by arranging environmental education programs, spreading general

environmental knowledge and by providing access to environmental information. This

empowering process may have positive implications on implementing pollution control

policies.

Society can play an effective role in manipulating the environmental performance of

pollution emitters. On one hand, they can impose pressure on polluters to adopt measures

to minimise pollution, while on the other hand, they can assist government regulators to

better enforce pollution control policies (Becker 1983; Murty 1995). Moreover, the

collective actions from the society can have an indirect effect on the level of pollution in

terms of market operation (i.e. consumption, investment and labour supply) (Pu 2003).

O’Connor (1998) emphasised that in developing countries consumers, investors and

workers by dint of their environmental preferences may create pressure on pollution

causing agents to improve their environmental performance.

Furthermore, World Bank (2000) has recognised that public participation plays a vital

role in pollution control whether the regulators are present or absent. In cases where

formal regulators are existent, communities adopt the political process to ensure the

enforcement of strict control measures. While, in cases where regulators are non-existent

or ineffective, Non-Government Organisations (NGOs) and community groups—

including water users’ association, religious institutions, social organisations, citizens’

movements and politicians—adopt informal regulation to put pressure on polluters to

conform to social and environmental norms (Afsah et al. 1996). However, legislative

provisions need to be designed to allow a significant degree of public involvement in the

pollution control process. Also, recognising the need of public involvement for pollution

control, emphasis should be given on designing appropriate environmental education

programs for the public and on ‘generating awareness through intensive publicity

campaigns using the press and electronic media, audio visual approaches, leaflets and

hoardings, as well as organising public programs for spreading the message effectively’

(Sharma 1997, p.311). Based on the evidence from Asia, Latin America and North

America, Afsah et al. (1996) suggests that neighbouring communities can have a

powerful influence on pollution discharge levels. Communities with higher income, more

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education and organisational capacity can more effectively influence polluters’

environmental performance and find ways to enforce environmental standards (Afsah et

al. 1996).

The local community and the voluntary groups may play an important role to monitor the

quality of river water on a regular basis. To promote such activity in the USA, the US-

EPA has developed detailed manuals for volunteer stream monitoring programs (US-EPA

1997). Besides, Ongley (2000) has suggested that in developing countries a new model of

decentralised community-based monitoring would be more effective than the traditional

model of centralised (regulator-based) monitoring system. For instance, community

groups can monitor and assess the health of river systems by using biological assessment

method, recording the abundance of macro invertebrates in a stream, rather than depend

on more conventional and expensive chemistry-focused approach (Ongley 2000; ANU

green 2006). Moreover, in the context of developing countries where monitoring is costly

and the budget of pollution control agencies’ are meagre, regulators can take pollution

control measures by responding to community complaints (Ongley 2000). However,

although complaints are a valuable source of low-cost information, regulations framed on

the basis of complaints may suffer from serious biases.

Public disclosure of information

This involves collecting information on the pollution emitters’ environmental

performance on a regular basis and disseminating that information to the public

(Blackman 2009b). This instrument encourages changes in polluters’ behaviour as it

discloses information about pollution among the general public. The information may

include pollution emission rate, ambient water quality conditions, damage costs,

abatement costs and performance ratings. According to some scholars, this instrument has

been characterised as the third wave in pollution control policy, after CAC and EI based

policy approaches (Hartwick and Olewiler 1998; Tietenberg 1998). A major advantage of

public disclosure is that it may operate as a useful mechanism for the regulators to

generate information on polluters’ performance. Here, the information can be used to

explore the possible opportunities for monitoring and control. Based on gathered

information, the authorities can set priorities and select more appropriate instruments.

Moreover, the collection of information is a positive sign indicating that the authorities

29

are growing more serious about the pollution and this can have important effects on

overall pollution management (Lopez et al. 2004).

The literature suggests that public disclosure can motivate polluters to cut emissions as it

generates pressure through external agents including: consumers who buy their products,

international certification bodies, various institutions who provide financial capital, the

employees of the enterprise, community groups, NGOs, industry associations, regulators,

legislators and the judiciary. In addition, literature also advocates that public disclosure

may inspire abatement by improving pollution causing agents’ internal information about

and attitudes towards pollution control (Tietenberg 1998; Blackman et al. 2004).

Blackman (2009b) postulates that applications of such instrument do not necessarily rely

on the regulators’ capacity for effective enforcement or on a clear set of environmental

regulations. Moreover, the necessary costs for administrative activities for data collection

and dissemination appear to be decreasing due to new information technologies.

Public disclosure is being considered as a means of overcoming constraints on

conventional environmental regulation in developing countries, where weak regulatory

institutions and incomplete regulations are an impediment in pollution control (World

Bank 2000; Dasgupta et al. 2007). Recent evidence from some developing countries also

suggests that public disclosure regarding the performance of the pollution causing agents

could be a powerful instrument for pollution control. For instance, Indonesia’s twin

public disclosure programs named (in Indonesian) PROKASIH and PROPER (translated

as Clean River Programs) are considered as classic examples in successfully causing

polluters to cut their emissions (in terms of BOD and COD) up to about 32 per cent

within ten years’ time (Afsah et al. 1996; Tietenberg 1998; Afsah et al. 2000; Lopez et al.

2004). Following the success of the programs in Indonesia, which was originally

launched in 1989, some other Asian countries like the Philippines (EcoWatch), China

(Greenwatch) and Vietnam (Green Bamboo) have also launched public disclosure

programs respectively in 1997, 1999 and 2002.

Voluntary agreements

These are usually non-legally binding contracts, either between industry (or any other

pollution causing agent) and the government or between industry and local community, in

30

which the pollution emitter in question volunteers to reduce its pollution by a certain

amount within a specified time (Lyon and Maxwell 2002; Blackman 2009b). The

objectives of voluntary environmental programs differ in industrialised countries from

those in the developing countries. In industrialised countries, regulators resort to

voluntary agreements to encourage polluters to comply with mandatory regulations or to

minimise emissions of pollutants for which no mandatory regulations formally exist. By

contrast, in developing countries, regulators generally use voluntary agreements to help

resolve extensive noncompliance with mandatory regulation (Blackman et al. 2006).

The main attraction of such an agreement to the industry in question is the public

endorsement that a successful pollution abatement program will bring in. On the other

hand, the agreement may be used by the government to put pressure on non-complying

industries which fail to keep their commitments—for instance by publishing information

pertaining to the results of inadequate abatement program of the industry (Murty et al.

1999). Blackman et al. (2006) verifies through a ‘game-theoretic model’ that when

voluntary agreements permit a significant increase in penalties, they may be prospective

in increasing investment in pollution abatement in developing countries, where the

regulatory capacity is relatively weak. However, like any other policy instruments, it has

some disadvantages as Lorenzen et al. (2007) argues that such agreements may increase

the risk of setting low goals of pollution abatement. They have pointed that voluntary

agreements may not be the most appropriate policy option to control the emissions of

hazardous substance, as the negotiated solutions and voluntary compliance may be

ineffective to set environmentally acceptable level of emission.

2.6 Comparison of alternative pollution control instruments

Based on the review of different pollution control instruments, Perman et al. (1999)

discuss a number of criteria that influence the selection of environmental policies. Those

criteria include dependability, information requirements, enforceability, long-run effects,

dynamic efficiency and flexibility. As a general reference, a comparison of selected

pollution control instruments is presented in Table 2.1 against those criteria and also

based on the discussions in previous sections.

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Table 2.1. Comparison of selected alternative pollution control instruments EI based instruments Criteria# CAC based

instruments Pollution charges

Marketable permits

Informal instruments

Dependability: to what extent can the instrument be relied upon to achieve the pollution target?

***

**

**

*

Information requirements: how much information does the instrument require that the control authority have, and what are the costs of acquiring it?

**

**

**

*

Enforceability: how much monitoring is required for the instrument to be effective, and can compliance be enforced?

**

***

***

**

Long-run effects: does the influence of the instrument strengthen, weaken or remain constant over time?

*

***

***

**

Dynamic efficiency: does the instrument create continual incentives to polluters for to reduce the extent of pollution?

*

**

**

*

Flexibility: is the instrument capable of being adapted quickly and cheaply with conditions or targets changing?

*

**

**

**

Note: Number of stars indicates how well the instruments meet the criteria. * stands for small extent; ** stands for medium extent; *** stands for great extent #Criteria adapted from Perman et al. (1999, p.223-224)

Bernstein (1997) suggests that each country should formulate clear and transparent

criteria to facilitate the selection of most appropriate instruments. For example, in

developing countries (like Bangladesh), where limited financial resources and weak

institutional capacity are unavoidable impediments, the two most important criteria could

be ‘cost-effectiveness and administrative feasibility’ (p. 183). Moreover, while selecting

the instruments for application, the policy makers also ‘need to take into account the

nature of the environmental problem and its causes, as well as practical, economic and

political realities’ (Bernstein 1997, p. 183). On the basis of the theories and the concepts

discussed in this chapter, detailed findings on the assessment on water quality of the

Buriganga River and the economic analysis on alternative policy options for pollution

control in this river are presented in the following chapters. These findings and the review

of relevant literature facilitate to develop an integrated pollution management framework

for this river.

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