The Pennsylvania State University

The Graduate School

Environmental Pollution Control

EXTENDED PRODUCER RESPONSIBILITY PROGRAM FOR

HOUSEHOLD HAZARDOUS WASTE MANAGEMENT AND HUMAN

HEALTH RISK ASSESSMENT IN DEVELOPING COUNTRIES:

ULAANBAATAR CITY, MONGOLIA

A Thesis in

Environmental Pollution Control

by

Temuulen Murun

© 2015 Temuulen Murun

Submitted in Partial Fulfillment

of the Requirements

for the Degree of

Master of Science

December 2015

 

 

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The thesis of Temuulen Murun was reviewed and approved* by the following:

Shirley Clark Associate Professor of Environmental Engineering Program Coordinator, Master`s of Science in Environmental Pollution

Control Thesis Advisor

Yen-Chih Chen Associate Professor of Environmental Engineering

Kathleen Raffaele Senior Science Advisor Office of Solid Waste and Emergency Response, U.S. EPA

*Signatures are on file in the Graduate School

 

 

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ABSTRACT

Over two decades ago, the European Union developed and started

implementing the Extended Producer Responsibility (EPR) policy to prevent household

hazardous waste (HHW) pollution and its environmental and human health risks.

Because of the increase in population, rapid urbanization, and the amount of importing

hazardous waste in developing countries, appropriate solid waste management and

disposal are needed especially for HHW treatment.

As a developing country, Mongolia does not have specific regulations and

treatment requirements for HHW. Unlike developed countries, the mandatory EPR

policy run by government will be necessary for Mongolia because small and family-

owned businesses are not willing to handle waste voluntarily due to the cost of

collection, recycling, and final disposal. Moreover, a Producer Responsibility

Organization (PRO) will face difficulties if advanced recycling fees and end-of-life

fees from consumers fund the organization, since the country`s GDP per capita in 2014

was only US $11,508.98 (Trading economics). Therefore, under the EPR program,

cooperation between a municipality and PROs is necessary and mutually beneficial,

especially in waste collection and transporting.

This is especially important because of the potential health risks resulting from

incineration of HHW. Due to open burning fluorescent bulbs in the central waste

facility, it is estimated that 69,637kg Hg/yr could be released into the atmosphere

directly. Human health risks from mercury exposure, using a traditional risk

characterization procedure, showed potential neurological concerns (hazard quotient:

14.42, exposure time: 12hr/d) for children who live within 500m of this municipal

waste incineration. In contrast, renal effects due to lead exposure (82.28kg Pb/yr) from

lead-acid vehicle battery recycling factories showed a low health risk for children. To

prevent this children`s neurological effect implementation of the mandatory EPR

program for HHW such as automobile batteries, fluorescent bulbs, and used tires is

necessary in developing countries because most of municipal solid waste facility are

designed poorly to capture emissions of hazardous substances.

 

 

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TABLE OF CONTENTS

List of Figures………………………………………………………….………….......v

List of Tables……………………………………………………………………...…..vi

Abbreviation………………………………………………………………………….vii

Chapter 1 INTRODUCTION………..………………………….…………………..…1

Chapter 2 HOUSEHOLD HAZARDOUS WASTE MANAGEMENT…………...….4

Background of Extended Producer Responsibility program……………...…...4

The EPR program implementation in developing countries……..…..………..9

Municipal solid waste and household hazardous waste management in

Ulaanbaatar city, Mongolia…………...………….…...……………………...15

Municipal solid waste……..……………………..…………………...15

Household hazardous waste…………………..…….………………..17

General EPR policy framework in developing countries: Ulaanbaatar city,

Mongolia……………………………...………………………………….…..19

Chapter 3 MERCURY RISK ASSESSMENT…………………….….......………….28

Mercury exposure from municipal solid waste incineration…………………28

Hazard identification…………………………………………………………35

Physical and chemical properties of elemental mercury……………..35

Toxicology: Neurological effects in children………………………...36

Risk characterization…………………………………………………………40

Chapter 4 LEAD RISK ASSESSMENT………………………………..………...….43

Lead exposure from recycling lead-acid car battery factors…………………43

Hazard identification…………………….……………………………….......49

Physical and chemical properties of lead…………..………………...49

Toxicology: Renal effects in children...………………….…………..51

Risk characterization………………….………………………………...…....53

Chapter 5 CONCLUSIONS…………………………….……………………..….….55

References……………………………………………………………………………58

 

 

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LIST OF FIGURES

Figure 1: Trends of e-waste generated in the U.S……………….…………………….2

Figure 2: Extended Producer Responsibility program………....…………………...…4

Figure 3: EPR program adoption…………………………………………………...…5

Figure 4: Recycled material components from e-waste………………………….…..11

Figure 5: E-waste management in Switzerland…...…………………………….……12

Figure 6: The recovery rate for some household wastes……………………………..15

Figure 7: Electronic waste composition by weight…….…………………………….18

Figure 8: EPR policy by product type………………………………………………..20

Figure 9: PROs and municipal participation in the EPR program (Mongolia)...….…25

Figure 10: EU Battery Directive (labels)…………………………………………….27

Figure 11: Compact FL driver………………………………………….…………….29

Figure 12: The former MSW facility…………………………..….……………...….32

Figure 13: The former MSW facility location and the city center…………………...33

Figure 14: The metallic mercury airborne exposure pathway…………….………….34

Figure 15: Dose and renal and neurological effects of mercury airborne

exposure (chronic)………………….………………………………………...38 Figure 16: Dose-response of elemental mercury vapor (inhalation)…………………39

Figure 17: The pollutant zone within the 500 m radius……………………………...40

Figure 18: Internal structure of a lead-acid battery………………………....………..44

Figure 19: The chemical reaction in a lead-acid battery………………………....…..45

Figure 20: Typical recycling process of used LABs in developing countries……….47

 

 

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LIST OF TABLES

Table 1: Urban waste generation by regions…………………………………………..1

Table 2: Factors to develop the EPR policy …………………...........…………….…..9

Table 3: The EU classification of e-waste…………….……...……………………...10

Table 4: Japanese waste management laws………………………………………….13

Table 5: Annual municipal waste generation in Ulaanbaatar city, 2012…………….16

Table 6: E-waste export and import…………………….……………………………20

Table 7: Energy saving from recycled materials………………………..……………26

Table 8: Mercury source in MSW………………………………...………………… 28

Table 9: Mercury containing lamps……………………………………………….....30

Table 10: Mercury emission from waste incineration………………………………..34

Table 11: The amount of mercury released into environment…………………….…33

Table 12: Physical and chemical properties of mercury……………………………..35

Table 13: NOAEL and LOAEL of animals and the human……………………….....37

Table 14: The mean wind velocity and average mixing height…………………...…41

Table 15: Consumption of lead for batteries (1993)…………………...…………….43

Table 16: Lead components in automobile batteries…….………………………..…45

Table 17: Lifetime of automobile batteries (1995)……...…………………………...46

Table 18: The annual amount of lead from LABs (Ulaanbaatar city)…….…………47

Table 19: Physical and chemical properties of lead………………………………….50

Table 20: LOAELs for neurological disorders……………………………………….51

Table 21: Lead nephrotoxicity in humans (Dose-Response)………...………………52

Table 23: Matrix application for the EPR policy framework……………..………….54

 

 

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ABBREVIATIONS

ADF Advanced disposal fees

ASTDR Agency for Toxic Substances and Disease Registry

AT Averaging time

BW Body weight

BC British Columbia

CDI Chronic daily intake

CFL Compact fluorescent lamps

CA Contaminant concentration in air

FL Fluorescent Lamp

EOL End-of-life

EPA Environmental Protection Agency

ED Exposure duration

EF Exposure frequency

ET Exposure time

EPR Extended Producer Responsibility

GASI General Agency of Specialized Inspection

GDP Gross domestic product

HQ Hazard quotient

HHW Household hazardous waste

IR Inhalation rate

IMERC Interstate Mercury Education and Reduction Clearinghouse

JICA Japan International Cooperation Agency

LAB Lead acid battery

LCD Liquid Crystal Display

LOAEL Lowest-Observed-Adverse-Effect Levels

MRL Minimum response level

MSW Municipal Solid Waste

NOAEL No-Observed-Adverse-Effect Levels

NGO Non-governmental organization

NPO Non-profit organization

 

 

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OERR Office of Emergency and Remedial Response

OECD Organization for Economic Co-operation and Development

PC Personal Computer

PCA Paint Care Association

PRO Producer Responsibility Organization

PbB Blood Lead

RfC Reference concentration

RoHS Restriction of Hazardous Substances

SENS Sustainability Expertise Network Solution

TPL Toxicology Profile for Lead

UB Ulaanbaatar

UF Uncertainty Factor

UNEP United Nations Environment Programme

WEEE Waste Electrical and Electronic Equipment

WHO World Health Organization

 

 

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Chapter 1

INTRODUCTION

Because of the rapid increase in population in developing countries,

appropriate solid waste disposal is needed. This will both protect residents` health and

upgrade a city’s sanitation and hygiene. For those cities that use incineration without

air control equipment as a final waste disposal option, air quality of the cities will be

polluted significantly. Thus solid waste management has become a crucial problem

especially in developing countries in recent years. According to East Asia and Pacific

Region of World Bank (What a Waste: Solid Waste Management in Asia, 1999),

“The urban areas of Asia produce about 760,000 tons of municipal solid waste per

day. In 2025, this figure will increase to 1.8 million tons of waste per day, or 5.2

million cubic meter per day.” This data shows how dramatically the solid waste

generation is increasing in cities in developing countries. Table 1 shows that waste

generation in urban areas will be doubled by 2025. (The World Bank, What a Waste:

A Global Review of Solid Waste Management, 2012)

Table 1. Urban waste generation by regions

Region Urban waste (2012) Urban waste (2025)

Africa 169,119 (t/d) 441,840 (t/d)

East Asia and Pacific 738,9958 (t/d) 1,865,379 (t/d)

OECD countries 1,566,286 (t/d) 1,742,417 (t/d)

Latin America and the

Caribbean region

437,545 (t/d) 728,492 (t/d)

Therefore, to decrease the environmental burden, increase living conditions, save

expenditures on unsophisticated disposal methods and resulting health problems, the

developing countries have been working on how to create efficient municipal solid

waste (MSW) treatments.

Household hazardous waste (HHW) treatment in developing countries is the

main issue in a municipal waste line because it contains potential hazardous

chemicals, which will need advanced treatment and most of the countries burn it

outside or dump it to open-sites. According to the U.S. EPA, HHW includes paints,

 

 

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consumer electronic products, batteries, fluorescent bulbs, pesticides, etc. All these

products can cause similar problems to industrial hazardous waste in the environment

if there is enough accumulation or inappropriate treatment for the waste. HHW has

been increasing in both developed and developing countries due to economic

development and rapid urbanization in cities. For example, according to Baeyens

(2010), “Non-rechargeable batteries contribute up to 90% of the total battery

consumption in Europe.” For the paint and related products, it also could contain

heavy metals such as cadmium, chromium, selenium, and lead. Paintcare (2012) states

“in the U.S., approximately 10% of paint sold in California is classified as “leftover”

paint, which equaled 6 million gallons in 2010, with approximately 70% being latex

paint.” Figure 1 shows the increase in electronic waste from consumer and industrial

usage in the U.S. (e-wasteregulation2, 2010). The same increasing pattern of personal

and office computer ownership and use that is one of the main HHWs at the end of

their life can be seen in developing countries.

Fig 1. Trends of e-waste generated in the U.S.

To prevent HHW pollution and human health effects, especially in developing

countries, the waste policy, regulations, and standards have to be systematic. Also it

should include responsibilities for consumers and producers as to how they will

handle the post-consumer waste. According to the OECD report, the most developed

and sustainable HHW management is the Extended Producer Responsibility (EPR)

program, which requires producers to take all the physical and financial

 

 

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responsibilities for their products` waste. However, in developing countries like

Mongolia, it is anticipated that it will need municipality participation in the EPR

policy for it to be effective.

To assess the risk of inappropriate disposal of HHW, risk assessment can be

used to analyze and characterize human and environmental health risks from toxic

substances and other stressors. Overall, risk assessment can be divided into two

groups: human health assessment and ecological risk analysis (U.S. EPA). In the EPA

guidance, to assess human health risk, several steps are needed: Planning and scoping

processes, Hazardous identification, Dose-Response, Exposure Assessment and Risk

Characterization.

This paper will focus on the potential and benefits from implementing the

Extended Producer Responsibility policy framework for HHW management in a

developing country, Mongolia, to identify human health risks from heavy metals in

waste by analyzing exposure and risk assessment.

 

 

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Chapter 2

HOUSEHOLD HAZARDOUS WASTE MANAGEMENT

Chapter 2 will introduce the key factors of developing the EPR policy, based

on the OECD manual for governments. Then the chapter analyzes the EPR program

implementation in developed regions/countries (the European Union, Germany,

Japan, Swiss, and British Columbia in Canada) to develop lessons that can be used in

developing countries to create their own EPR programs. Municipal solid and

household hazardous waste management, and regulations in Ulaanbaatar (UB) city,

Mongolia also are introduced in this chapter. Finally, the chapter will discuss EPR

policy implementation in the city and its general framework, legislation, and

operational management.

Background Information of Extended Producer Responsibility program

According to a guideline manual for governments on extended producer

responsibility published by the OECD in 2001, a definition of EPR is “an

environmental policy approach in which a producer`s responsibility for a product is

extended to the post-consumer stage of a product`s life cycle.” In other words, the

EPR program is the most sustainable waste management policy because it states a

responsibility for producers to handle and treat the post-consumer waste of their

products to prevent environmental pollution and human health risks due to the

hazardous waste and toxic substances. Upstream and downstream waste management

in the EPR policy is shown in Figure 2 (McKerlie, Knight, and Thorpe, 2006).

Fig 2. Extended Producer Responsibility program

Manufacture    

Retail/Sale  

Waste  collection  

Trasportation  

Recycling/Reuse  

Recycled  and  raw  materials  

 

 

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Is the Extended Producer Responsibility program necessary?

Due to rapid urbanization especially in developing countries, the amount of

solid waste and household hazardous waste has been increasing; however, municipal

waste management and treatment capacity are not able to handle the waste. The

municipal revenue from taxpayers is not enough because of increased waste loading;

therefore, it needs to be funded by outside investment/non-tax revenue. Thus, to meet

the demand, there will be a need for a new waste management policy to reduce the

municipality’s burden physically and financially (OECD, 2001). In 1994, the OECD

began working and developing the EPR framework and the member countries

initiated implementation of the approaches in their national level policy (OECD,

2014). Since then, the number of waste and environmental policies that have adopted

the EPR program has been increasing around the world. Figure 3 shows an increasing

trend of EPR program adoption (OECD, 2014).

Fig 3. EPR program adoption

A Guidance Manual for Governments on Extended Producer Responsibility

In the manual (OECD, 2001), the governmental EPR framework and

individual participant`s responsibilities, cost and benefits are written in detail.

The main concerns that the EPR program has to answer are increasing collection and

recycling rates, producer`s financial responsibility, and improvement of product

design to consider the environmental impact. Further broader categories for designing

an EPR framework would include (OECD, 2014):

• Product take-back process:

 

 

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This program requires manufacturers to take back their products after usage and it can

be through retailers or contractors that are paid by producers. Sometimes, to increase

the collection and recycling rate, different kinds of incentives can be given to

consumers for bringing back the post-consumer products (OECD, 2014). If the

producers have to meet targets for collection and recycling, the operation will be more

efficient in order to meet the target.

• Market-based instrument:

The EPR program should be economically beneficial to each party: private industry,

communities, and government authorities. For instance, deposit-refund programs,

advanced disposal fees (ADF), and natural material taxes can be incentives for

stakeholders.

1. Deposit-refund programs:

In the OECD guideline (2001), the deposit-refund system is defined as a

program where “a payment (deposit) is made when the product is purchased

and is fully or partially refunded when the product is returned to a dealer or

specialized treatment facility”. Retailers return the deposit when they receive

the same brand which they sell. It seems that a higher deposit price encourages

a higher return rate (OECD, 1994).

2. Advanced disposal fees (ADF):

According to the OECD guidance (2001), “an advanced disposal fee would

be a fee levied on certain products or product groups based on estimated

costs of collection and treatment methods“. Government, a private industry, or

a Producer Responsibility Organization (PRO) could collect the fee; however,

it should be clarified in the planning of the EPR program. Even though some

OECD member countries have ADFs at the point of sale, they have a system

to return a portion of the fee due to reduction of recycling fees for the products

(OECD, 2001).

3. Virgin material taxes:

The goal of the material tax is to increase recycled material usage and

reducing the virgin material proportion in the products. Also, a special tax can

be levied on potential hazards and toxic materials to encourage reducing

environmental pollution (OECD, 2001). The revenue from these taxes

 

 

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would be used for waste collection, treatment, and the cost for managing the

system.

• Regulations and standards:

The EPR program can be mandatory or voluntary based on the country`s economic

growth, infrastructure, and socio-cultural situation, including residents` education and

public support. Voluntary programs can be established between industries and public

authorities or private sectors and non-governmental organizations (NGOs). Some

countries in the EU have a target percentage for several post-consumer waste

categories to be recycled and minimum recycled content requirements for paper

products, beverage bottles, and plastic containers (OECD, 2001).

• Information-based instruments:

Public awareness for environmental policies and human health effects due to

pollutants is important to implement the EPR program successfully on the national

level. Therefore, to raise this awareness, producers’ annual reports of recycled rates,

clear labels for the products having a potential hazard to humans and the environment,

waste separation reports, communication between producers and consumers, and the

rate of second hand or recycled materials usage in products can be strong

measurements for producers to comply with EPR policy.

The OECD Global forum on environment (2014) stated several important

factors are necessary to develop the EPR program. Based on different factors/targets,

the program`s outcome will be various.

• Type of product

The program covers only one product such as televisions and car batteries or a group

of products like any electronic waste or used vehicle oil. According to the OECD

(2014), small electronic waste is easily covered by an EPR program.

• Voluntary or Mandatory

Whether the EPR program can be voluntary or mandatory is based on agreement

between government authorities and private sectors. If the producers are willing to

participate the program, it may not be necessary to legislate the EPR policy. A

voluntary EPR program can be referred to as a Stewardship program (OECD, 2014).

Nicol and Thompson (2007) stated the Product Stewardship program is a shared

responsibility system; therefore, producers do not have financial and physical

 

 

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responsibility of post-consumer waste for disposal and treatment. They argue that the

policy fails to prevent pollution from waste and reduce consumption.

• Operation of waste management

The manufacturer could handle the collection and recycling process individually or

producers could establish a PRO that is funded by member producers based on their

products` recycling methods, hazardous substance content, and weight. The

organization acts on behalf of their member manufacturers; therefore, PROs could

collect and recycle post-consumer waste (OECD, 2014). More than 260 PROs were

established in Europe in between 1998 and 2007 (Mayers, 2007).

• Financial or Organizational responsibility

At the global forum (OECD, 2014), it stated the producers could select the financial

or organizational responsibility for their products` waste management.

1. Financial responsibility: Producers or PROs pay municipalities annually

for handling the post-consumer waste. Recycling and collection could be

treated by specific contractors or municipalities.

2. Organizational responsibility: Producers or PROs finance and operate

waste management and treatment or contract with recyclers directly.

• Responsibility among stakeholders:

The goal of the EPR program is to shift the responsibility for waste management from

municipalities to manufacturers. However, in most countries, municipalities are still

in charge of the collection and sorting due to their capability. The municipality has

sufficient storage and infrastructure to handle waste. Once post-consumer waste is

sorted, then it is transferred to the manufacturers for recycling treatment. Different

countries have various responsibilities for stakeholders; however, even for consumers,

the responsibility should be clarified. For instance, in Japan, final consumers are in

charge of separating their waste and paying a recycling fee for electronic wastes:

television, washing machine, etc. (OECD, 2014).

• Cost for full operation:

The main concern is how to calculate the full cost of end-of-life coverage and how

much the producers should cover fees for handling waste. The OECD manual (2001)

discusses the full cost including collection, sorting, recycling, awareness campaigns,

public advertisement, monitoring, and reporting. For the small businesses and

 

 

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importers, the full cost of waste management and operation will be a burden, so it

needs to discussed and defined clearly in the EPR context.

• Free riders, orphan and existing products

In the guideline, free riders are defined as “the actors in EPR system who do not pay

for the benefits they receive” (OECD, 2001). If the product`s producer has gone

bankrupt or does not exist anymore, the product is an orphan product. Existing

products also were produced before the EPR policy was implemented. All these

products obtain advantages that are not regulated under the EPR program, which can

make a negative impact on the system. According to the OECD manual (2001), peer

pressure, monitoring, reporting, and sanctions can be tools for avoiding free riders.

Moreover, government authorities would make standards and rules for minimizing

and eliminating them. Management of the orphan and existing products depends on

cost of waste treatment, recycling methods, and the numbers of products that are in

the market (OECD, 2001).

Based on the above several factors, examples of different types of approaches

for development of the EPR program are shown in Table 2 (OECD, 2001).

Table 2. Factors to develop the EPR policy

Factors Take-back program 1. Mandatory

2. Voluntary Legislations 1. Target percentage for recycling

2. Disposal bans 3. Prohibition for hazardous

substances 4. Usage of recycled material

Economic approaches 1. Deposit-refund 2. Advance recycling fees 3. End of life fees 4. Natural material taxes

The EPR program implementation in developed countries  

• European Union and Germany

Under the EPR policy, the European Union`s waste management has been

leading the world in sustainable management and reduced environmental burden. The

main legislations are: End-of-life Vehicles Directive, Packaging and Packaging

Waste Directive, Waste Electrical and Electronic Equipment (WEEE) Directive, and

 

 

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Restriction of Hazardous Substances (RoHS) (McKerlie, Knight, and Thorpe, 2006).

According to the WEEE Directive (2003), member countries` manufacturers have to

take responsibility for all costs of collection, sorting, recycling, reusing, and disposal

stages for the products` post-consumer waste. This system leads to innovative

recyclable designs for the products in early manufacturing stages (McKerlie et al,

2006). Moreover, the Directive reduces the amount of e-waste going to landfills and

incinerators by promoting the recycled material proportion and the target-recycling

rate. Overall, European countries have around 114 PROs for packaging waste

management, 17 PROs for batteries` end-of-life disposal, and 129 WEEE PROs

(Mayers, 2007), which indicates the PRO system is well developed financially and

operationally. Table 3 illustrates the EU classification of WEEE and its labeling

information (Khetriwal, Kraeuchi and Widmer, 2009; Nnorom and Osibajo, 2008).

Table 3. The EU classification of e-waste

WEEE Category Special label

1. Large home appliances Large HH

2. Small home appliances Small HH

3. IT and telecommunication equipment ICT

4. Consumer equipment CE

5. Lighting equipment Lighting

6. Electronic tools E&E tools

7. Toys and sports equipment Toys

8. Medical devices Medical equipment

9. Monitoring and control instruments M&C

10. Automatic dispensers Dispensers

The RoHS Directive was initiated in 2002 and began in July 2006. It limits six

potential hazardous substances: “lead, mercury, cadmium, hexavalent chromium, two

flame retardants added to plastics: polybrominated diphenyl ethers and

polybrominated biphenyls” in order to prevent their use in electronic products ranging

from small electric toys to washing machines (McKerlie et al, 2006; Kahhat et al,

2008; Sthiannopkao and Wong, 2013). The European Directives have had a

significant impact on the international electronic market by requiring producers to

 

 

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redesign products with fewer hazardous materials and in preparation for easy

recycling and dismantling processes.

The Green Dot program for packaging waste is the first and largest EPR

program in Germany and the EU. According to the program, every producer of

packaged products is required to establish their own take-back program or participate

in PRO: Duales System Deutschland (McKerlie et al, 2006). Member manufacturers

have to pay an annual fee to use the eco-label that indicates that there is a free

collection program. This program and labeling make it easy to notice and convenient

for consumers. The fees are different based on the material and weight of products;

for instance, tin or paper products have low fees because they are made of highly

recyclable materials (McKerlie et al, 2006). In total, Germany has eight PROs for

packaging waste, three PROs for end-of-life of batteries, and twenty PROs for WEEE

management (Mayers, 2007). McKerlie et al (2006) concluded that the Green Dot

program for packaging waste has resulted in more efficient recycling package design

and waste reduction to final disposal.

• Switzerland

The European Union`s WEEE Directive was legislated in 2003; however, the

Swiss have had take-back e-waste legislation since 1998 (Khetriwal et al, 2009). The

largest two PROs: SWICO Recycling Guarantee and SENS Swiss Foundation for

Waste Management collect and dispose of waste for their member producers and

importers. Figure 4 shows the recycled material from WEEE under the two PROs:

SENS and SWICO (Widmer et al, 2005).

Fig 4. Recycled material components from e-waste

 

 

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In 2007, four PROs in Switzerland that are all non profit organizations (NPOs) treat

most of the e-waste from households to offices, including the categories 1-5, 6, and 7

of the EU classification of WEEE (Table 2).

Consumers are obligated to return used products to retailers or collectors and

to pay the advanced recycling fee (ARF) for newly purchased products. Therefore,

there should be less illegal electronic waste dumping because no disposal fees need to

be paid in the post-consumer stage. According to Khetriwal et al (2009), the recycling

fee paid by consumers is calculated by two methods: product price index (SWICO)

and six different fee categories (SENS). Under the Swiss law, retailers should inform

consumers that the price includes the ARF; moreover, PROs recommend putting a

visible recycling fee to raise awareness about recycling in the end-of-life stage. The

ARF is collected from consumers, manufacturers, and importers to PROs to fund

collection, transport, storage, dismantling and recycling of disposed products. Figure

5 shows the electronic waste EPR system in Switzerland (Khetriwal et al, 2009).

Fig 5. E-waste management in Switzerland

• Japan:

In 1998, the Japanese government enacted the Home Appliance Recycling

Law and implemented it nationwide in 2001. Under the law, producers and importers

are responsible for managing the recycling of four types of home appliance:

televisions, refrigerators, washing machines, and air conditioners (Kahhat et al, 2008).

The last consumers are obligated to pay a part of the fee for recycling, transportation

and to sort the appliance from municipal waste. The end-of-life fee is based on the

 

 

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type of products and is around US $23-44 for recycling processes and US $4-18 for

transportation (Ogushi and Kandlikar, 2007). The main disadvantage of the end-of-

life (EOL) fee paid in the disposal stage is that it could increase illegal dumping.

Around 2 percent of EOL home appliances (172,000 units) were illegally discharged

in 2004 and it is higher than illegal dumping in 2001 (Ogushi and Kandlikar, 2007).

Mostly municipalities are in charge of collecting and transferring the end-of-life

electronic products, but the individual sector`s role in collecting e-waste has been

increasing. The Association for Electric Home Appliances of Japan handles orphan

products (Nnorom and Osibanjo, 2008).

In 2001, Japan legislated the Law for Promotion of Effective Resource

Utilization to cover used computers from the business sector and individual

consumers. A Personal Computer (PC) 3R (Reduce, Reuse, Recycle) Promotion

center was established by manufacturers to collect and recycle PCs sold after Oct

2003. PCs covered by the program have a “PC Recycling Mark” to indicate a non-

refundable recycling fee and the payment is paid at the point of sale. However, for the

computers sold before Oct 2003, the final consumers need to pay recycling and

collection fees of products (Kahhat et al, 2008). In 2004, at the G8 Summit, Japan

introduced the 3R program to the international stage to promote sustainable waste

management (Sthiannopkao and Wong, 2013). The program prioritizes reducing the

waste generation at source points, reusing valuable materials from products, and

recycling materials from waste for different purposes.

The End of Life Vehicle Recycling Law requires the customer to pay a

recycling fee at the time of the new car purchase (Ogushi and Kandlikar, 2007) to

prevent illegal disposal. Under the EPR policy, the success of the Japanese recycling

process is due to environmentally focused development in the manufacturing stage

and central recycling factories around the country. Table 4 shows the main waste

management laws in Japan (Ogushi and Kandlikar, 2007).

Table 4. Japanese waste management laws

Law Content Year enforced

Recovery rate (Fiscal yr)

Fundamental Law for Establishing a Sound Material Cycle Society

Products are need to be recycled

2001

 

 

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Waste Management Law Required 3R program in 10 industries and 96 products

2001

Law for Promotion of Effective Utilization of Resources

Desktop PCs Laptop PCs Cathode ray tube Liquid crystal display Ni-Cd batteries Ni-MH batteries Li ion rechargeable batteries Small lead batteries

2001 76 (2005) 54 (2005) 78 (2005) 68 (2005) 74 (2004) 77 (2004) 55 (2004) 50 (2004)

Containers and Packaging Recycling Law

Cans, bottles, cartons and boxes

1997

Home Appliance Recycling Law

Air conditioners Televisions Refrigerators and freezers Washing machines

2001 84 (2005) 77 (2005) 66 (2005) 75 (2005)

Food recycling Law Food waste 2001 45 (2004) Construction Material Recycling Law

Concrete Wood Asphalt

2002 98 (2002) 89 (2002) 99 (2002)

End of Life Vehicle (EOL) Recycling Law

EOL vehicles 2004 57-68 (2005)

• British Columbia, Canada:

The British Columbia (BC) government started developing a municipal waste

management plan in 1989 and the strategy was approved in 1995. The main goal of

the plan was to prevent pollution from household hazardous waste such as paints,

pesticides, and batteries (Driedger, 2002). In 1994, the Waste Reduction

Commissioner required manufacturers and producers to take back and recycle the

eight HHW; therefore, the producers of paints and related products established the BC

Paint Care Association (PCA) to treat and dispose of used paints. The manufacturers

need to pay an eco-fee (C$0.50/4L container and C$0.10/aerosol container) to the

PCA for the collection, recycling, and final disposal (Driedger, 2002). Similar to the

paint treatment system, the EPR program for HHW such as hazardous

pharmaceuticals, drugs, pesticides, and automobile oil products has been expanding

significantly. Due to the successful closed-loop waste management system in BC

Canada, the recovery rate on household waste is more than 70 percent. Figure 6

 

 

15  

illustrates the recovery rate for some common household wastes: containers, car

batteries, tires etc. (McKerlie, 2006).

Fig 6. The recovery rate for some household wastes

The unique feature for the membership fee system in the Paint and Product Care

Association is that the fee is based on producer`s market share capacity (Driedger,

2002). This can be a useful approach for the small manufacturers and importers to

collect a fee, plus there is a potential reduction in the number of free riders and orphan

products. The eco-fee paid by consumers is not a federal tax in Canada, so the fee is a

part of the price charged by the manufacturers and retained by the manufacturers.

This requires the manufacturers to highlight the reasons for and uses of the eco-fee.

Municipal solid waste and household hazardous waste management and

treatment in Ulaanbaatar city, Mongolia

Municipal solid waste Between 2006 and 2010, a new landfill plant funded by Japan International

Cooperation Agency (JICA) was built in UB city to treat municipal solid waste. The

city has five waste treatment centers and population is approximately 1372 (National

Plan for Waste Management Improvement, 2014 and Statistics Department of

Ulaanbaatar city, 2014). However, open dumping of waste from construction and

open burning of household solid wastes are common nationwide. General

consumption also has been increased due to mining sector development and foreign

investment; for example, in 2011 the percent from the previous year was increased 8.9

96%   91%  75%  

99%   96%  

0%  

20%  

40%  

60%  

80%  

100%  

120%  

beer  containers  wine  and  spirit  containers    

non-­‐alcohol  beverage  containers  

car  batteries     tires  

Recovery  rate    

 

 

16  

percent for imported chemical substances, 28.6 percent for office equipment and

electronic products, 28.1 percent for vehicle batteries and tires, and 71.8 percent for

secondhand vehicles (National Plan for Waste Management Improvement, 2014). Due

to an overwhelmed waste generation and treatment system in UB city, in 2012, the

Mongolian government passed a new waste law that includes municipal solid waste,

industrial waste, and hazardous waste (Waste Law, 2012). According to the law, each

city`s authority is responsible for collecting only municipal waste that is generated

within its territory and treating the disposed waste in a centralized waste facility. For

industrial waste and hazardous waste, manufacturers are required by the law to

transport waste to the centralized treatment facility and to pay a fee for the final

disposal. The fee depends on the weight and hazardous components of the waste

(Waste Law, 2012). The budget for MSW treatment consists of mostly tax revenue

and the disposal service fee from industries. Therefore, there has been a huge burden

on the municipality to handle both consumer and industrial waste given the small

budget for waste treatment.

According to the final JICA report for waste management improvement in UB

city (2012), the MSW facility, the preferred option for household waste disposal, has

high emissions of pollutant gases, and increased risks for both human health and

environment. The amount of waste generated in the city is shown in following Table 5

(Final JICA report, 2012).

Table 5. Annual municipal waste generation in Ulaanbaatar city, 2012

Household waste generation (ton/day) 1096.6

Waste from business activities (ton/day) 397.9

Waste from public cleaning (ton/day) 31.8

Total (ton/day) 1526.3

Due to the rapid movement of people from the countryside to the city and the

construction boom in the capital city, the annual amount of waste likely has increased

since 2012. For instance, in 2009 the amount of annual solid waste was 820,000 tons

at the national level; in 2013 the amount was up to 2.4 million tons (National Plan for

Waste Management Improvement, 2014). Under the National Plan (2014), the waste

improvement strategy is going to be implemented in two phases.

 

 

17  

1. The first phase (2014-2017):

To develop the legal environment for waste reduction, management improvement,

infrastructural and financial enhancement, and community participation in waste

reduction activity

2. The second phase (2018-2022)

To strengthen environmental remediation, green products usage, and systematic waste

management

Household hazardous waste

According to the Waste Law (2012), hazardous waste is the waste that has

negative impacts on humans, animals, plants, and their descendants, is hazardous,

corrosive, flammable, explosive, contagious, radioactive, pollutes the environment

and can be solid, liquid, and gas state. In the law, the hazardous waste from industries

has to be separated from MSW and treated in a facility that has advanced technologies

to reduce the hazard. Once treated, the waste can be transferred in landfills (Waste

Law, 2012). However, the law does not legislate what is household hazardous waste

(HHW) and whether it is included in the definition of hazardous waste or not. Since

there is no legislation on HHW, the city has no segregation and sorting programs for

any HHW from municipal solid waste. Therefore, the UB city municipality has no

specific data on the annual HHW generation.

In the U.S., according to the EPA, regarding wastes generated by households

and household-like areas, “Leftover products that contain corrosive, toxic, ignitable,

or reactive ingredients are considered to be household hazardous waste.” Therefore,

HHW could be consumer electronic waste, old solvents, oil-based paints, batteries,

pesticides, fertilizers, or fluorescent lamps. In addition, medical sharps,

pharmaceuticals and poisons from households can contain toxic substances and heavy

metals (EPA, Non-hazardous waste report). Unlike Mongolia, the EPA specifies these

wastes as HHW.

Under the Waste Law (2012), there is no specific recycling target for HHW

and municipal solid waste and it states recycling and reusing processes are

recommended if producers are willing to do it. It is not a mandatory program;

 

 

18  

therefore, manufacturers and importers have no incentive to recycle their products`

post-consumer waste.

WEEE contains potential hazardous heavy metals such as lead, mercury,

cadmium, and nickel; reusing and recycling the rare metals from e-waste would

reduce energy requirements for natural resource extraction and processing of crude

materials. According to Widmer et al (2005), 40 percent of consumer personal

computers end up in landfill and heavy metals from e-waste are 70 percent of the total

heavy metals in the U.S. landfills. Figure 7 illustrates the material compositions in e-

waste (Widmer et al, 2005).

Fig 7. Electronic waste compositions by weight

As shown in Figure 5, e-waste weight contains more than 50 percent metals, including

both ferrous and non-ferrous metals that could be recycled. The final JICA report

(2012) mentioned there are a few small e-waste recycling businesses in the city;

however, the recovery rate is only about 20 percent of e-waste generated. For

example, sometime individuals usually go to a computer shopping center to sell

broken personal computers or to purchase secondhand computers or recycled one.

In 2002, a new guideline for the medical waste separation, collection, storage,

transportation and final disposal was legislated by the Ministry of Health and the

Ministry of Environment of Mongolia. Under the law, the specific medical waste

includes blood from humans and animals during medical treatment, medicine and

related products, bioactive and radioactive products, syringes, and medical equipment

(Enactment 249/201, Ministers of the Ministry of Health and the Ministry of

Environment, 2002). The enactment (249/201, 2002) states the specific medical waste

 

 

19  

from national hospitals funded by the government is collected, stored, and treated by

tax revenue. However, waste from medicine and related products` in individual

households are not included in the law; therefore, pharmaceutical and medical waste

is mixed with municipal solid waste. The list below shows the situation and resulting

problems of HHW management in UB city:

• No specific HHW management and policy exists in UB city

• HHW is treated in the municipal solid waste facility

• There is no advanced technology in MSW treatment

• The primary treatment is incineration without controls

• After incomplete incineration, the waste mostly goes into landfills

Mongolia has specific laws, rules, and standards for industrial hazardous

waste management; however, there has been a lack of regulation for HHW and less

monitoring for industrial hazardous waste treatment. Due to not having advanced air

control technologies and landfill leachate systems in MSW treatment, the potential

hazard for human health is increased dramatically.

General EPR policy framework in developing countries:

Ulaanbaatar city, Mongolia

Due to a variety of including cost benefit, restrictive regulation, and a lack of

recycling infrastructure, most of the waste such as e-waste, used tires and plastic

waste is transported from developed countries to China, India, Pakistan, and Nigeria

to final disposal (Schiannopkao and Wong, 2013). These waste-importing countries

may not have systematic waste regulations or advanced treatments; therefore,

treatment in these countries likely results in increased industrial pollution. For

example, “open burning of coated fires, heating of printed circuit boards to remove IC

chips, and acid baths for extracting gold” are well known in China (Kojima et al,

2009). Table 6 shows the amount of e-waste is exported and imported (Schiannopkao

and Wong, 2013).

Table 6. E-waste export and import

Countries

or region

Household waste

(millions of tons)

Exported

(millions of tons)

Imported

(millions of tons)

 

 

20  

USA 8.4 2.3 -

EU 25 8.9 1.6 -

Japan 4.0 0.59 -

China 5.7 - 2.6

India 0.66 - 0.97

As seen in Table 6, there is no imported waste in developed countries; thus, importing

countries need to develop hazardous waste regulations and standards for preventing

human health risks caused by e-waste pollution.

• Legislation and restriction

Since Mongolia has no legislations/regulations or classifications on HHW and

less monitoring of industrial hazardous waste treatment, the government needs to

initiate classifying, regulating, and developing the EPR policy for hazardous waste

collection and treatment from consumers and manufacturers. Kaffine and O`Reilly

(2013) concluded the EPR program is suitable for consumers` e-waste, automobile

batteries, packaging, and tires. Figure 8 illustrates the products are covered by the

EPR policy in globally (Kaffine and O`Reilly, 2013).

Fig 8. EPR policy by product types

The “other” (20%) includes used oil, paints and related products, and pesticides

(Kaffine and O`Reilly, 2013). Mongolia could implement the EPR policy for e-waste

and automobile batteries, since both e-waste and imported vehicle parts such as

batteries has been increasing. In terms of developing the EPR program for hazardous

waste, the scope of the program, a consensus goal and target, objectives and

35%  

11%  17%  

17%  

20%   E-­‐waste  

Vehicle  batteries  

Tires  

Packaging  

Other  

 

 

21  

responsibilities must be determined clearly and understandably for each actor (OECD

guideline, 2001). Otherwise, there will be misunderstandings between government

agencies and producers; for instance, Brazil has faced many difficulties: less

supportive legislations and confusion on prohibition of imported tires to implement

the EPR policy for used tires (Milanes and Buhrs, 2009).

After the EU`s countries, Japan, Canada, and Switzerland successfully

implemented the EPR program for their waste management, some developing

countries are trying to embrace this policy. One of the biggest recycling countries,

China, has legislated “the management of waste household electrical and electronic

products recycling and disposal” with its content being both mandatory recycling of

WEEE: TV sets, refrigerators, washing machines, air conditioners, and personal

computers under the EPR policy, and certifying second hand products and recyclers

(Kojima, Yoshida, and Sasaki, 2009, Hicks, Dietmar and Eugster, 2005). Another

effective legislation in China is “management measure for the prevention of pollution

from electronic products.” The law restricts toxic substances from products, requires

labeling for hazardous materials, and promotes green products. (Hicks et al, 2005).

These restrictions are established in the EU RoHS and the WEEE Directive in terms

of hazardous substance restrictions and e-waste disposal. The following are the six

toxic chemicals for which China restricted usage in electronic products: “Lead,

mercury, chromium IV, cadmium, polybrominated biphenyls and polybrominated

diphenyl ethers” (Hicks et al, 2005).

Not only China but also several countries such as Japan, Taiwan, South Korea, and

Brazil developed their EPR policy based on the EU WEEE Directive and RoHS;

therefore, these regulations could be valuable information and an important guideline

for Mongolia to implement the EPR policy at national level. Most electronic products

and automobile related products or second hand vehicles are imported to Mongolia

from China and Japan, respectively. Therefore, under the hazardous component

restriction of an EPR program, imported appliances are easy to monitor, control, and

ban through the custom office. Moreover, a tax increase on imported products could

be established based on the proportion of hazardous substances and recyclable content

to encourage usage of less environmental impactful products.

• Mandatory EPR program

 

 

22  

Similar to other developing countries, Mongolia has many small importers and

family businesses, plus a few big importing companies; thus it can be seen the market

is not mature and most producers and importers likely will not be willing to comply

with a voluntary EPR policy due to the cost for collection, recycling, and final

disposal. Therefore, there should be the government involvement in the EPR program

as it relates to electronic product and automobile battery importers. According to the

OECD seminar on EPR in France in 2001, under the mandatory program, the free

rider problem is going to decrease, and higher collection and recycling rates are more

achievable compared to the voluntary program. Moreover setting a collection target is

an effective way to meeting the target; for example, in Belgium this way has

increased collector`s responsibility for batteries under the mandatory EPR program

(OECD seminar, 2001). In the voluntary EPR policy, there have been free rider issues

caused by non-participating producers and the competitive disadvantages for member

companies compared to the non-participants. Furthermore, several benefits of the

mandatory EPR program run by the government (OECD guideline, 2001, Nnorom

and Osibanjo, 2008) include the following:

• Eliminating hazardous substances and improving Designed for the

Environment products

• Clarifying the producer`s financial responsibility

• Implementing EPR supportive policies

• Raising awareness of the policy

• Establishing mandatory targets for the collection and recycling to promote

technological innovations

It could be both more flexible and time consuming for waste compliance and

permission if the EPR program is mandatory because city authorities and different

government agencies have a consensus national regulation and rule that are easily

spread throughout the country. This leads to less confusion between the government

and producers.

Mayers (2007) mentioned that if there are no targets and specified treatment

requirements, the policy does not send a direct and clear signal to manufacturers. For

example, it should clarify in the WEEE Directive what kind of separation treatment

would need to be used if LCD backlights contain mercury (Mayers, 2007). Therefore,

 

 

23  

under the mandatory EPR program, the government could provide general treatment

requirements, standards, and specific methods for recycling and final disposal in order

to increase efficiency and reduce company investment in treatment. A guidance

document or manual can be discussed and developed between producers and

government agencies to satisfy both party`s demands.

• Producer Responsibility Organization:

In terms of post-consumer waste treatment, individual producers and

importers are not able to handle all the required processing in Mongolia due to a lack

of technical equipment, qualified employees, and an efficient collection system. The

same situation is occurring in China and other developing countries, where individual

collection and treatment systems are not feasible (Nnorom and Osibanjo, 2008).

Therefore, PROs become an integral part of the EPR program in these locations

because they can address these limitations. Also, PROs can treat a broader type of

products than individual manufacturers, especially when those individual

manufacturers/processors are small family businesses. Finally, the PRO can save time

and reduce the cost of recycling and final disposal (Spicer and Johnson, 2004).

Therefore, the EPR policy would require all the actors to join a PRO that is funded by

a member companies` annual fee. The fee should be based on their market share,

products` size, weight, and hazardous substances, because a flat fee does not promote

waste reduction in general and a gap between small businesses and larger group

companies is expanding in Mongolia. Further, the flat fee penalizes the companies

that improved their products in relation to their environmental impact. These

producers in a PRO are not going to receive any financial benefits from the EPR

program (McKerlie et al, 2006).

In developed countries such as Japan, Germany, and Switzerland, most of

PRO activities are fully dependent on participating company`s funds and recycling

fees from consumers; however, in Mongolia, this system may not work successfully.

The recycling fees paid by consumers are a particular challenge. Consumers are

unlikely to accept the rule since the country`s GDP per capita (US $11,508.98 in

2014) is much lower than the European countries (GDP per capita US $32,789 in Dec

2014) (Trading economics). Therefore, in a developing country like Mongolia, using

the advanced recycling fee and the end-of-life fee from consumers could be

 

 

24  

challenging if they are the only revenue sources to fund a PRO. Moreover, e-waste

may be disposed to informal recycling factories since it has valuable metals and is

useful for reuse in secondhand electronic products. For the above the reasons, the

PRO operation requires municipal participation, especially in waste collection and

transportation, because the municipality already has the collection operation and the

PRO likely does not have the funding to establish a second collection and transport

system. Since UB city`s municipality has waste storage locations, developed

transportation and collection system, and waste return locations, incorporating the

municipality into the EPR program will reduce costs for building infrastructure. In

fact, in some countries such as the United Kingdom, China, and Taiwan, a local

municipality is responsible for the collection, separation, and transportation of HHW

(He et al., 2006, Nnorom and Osibanjo, 2008).

Further, the Ulaanbaatar city authority could allow PROs to use disposal

facilities and collection centers for recycling, which gives producers and importers a

feasible management location for waste treatment. Therefore, unlike other

developing country`s cases, PROs (producer and importers) would be responsible for

only recycling and final disposal. For the recycling processes, PROs would need to

cooperate with informal small recycling businesses to develop efficient waste

treatment. This is because the secondhand products especially used tires, vehicles, and

personal computers are still valuable in the market; therefore, many small family-

owned companies have been operating in the city. These small businesses must be

included in the PRO because the technologies used in this informal recycling sector

are not environmentally friendly and not meeting with the country`s standards. The

PROs could provide the International Organization and Standardization standard

technologies and advanced equipment for recycling. In terms of the waste handling

processes, PROs should not have brand-specific rules in a developing country like

Mongolia because the market itself is small comparing with developed countries and

there is not enough fund for each brand`s post-consumer waste, especially personal

computers, televisions, and refrigerators.

Under the full industry-responsible EPR waste management program, there is

a direct feedback loop from the recycling processes to the manufacturing or product

design phases. This could allow designers to change product`s recyclability and to use

 

 

25  

alternative materials. On the other hand, in the PRO system, especially one not

focused on specific brands, there is no direct feedback loop and feedback from PROs

may not be an effective influence on a product design (OECD, 2001, Spicer and

Johnson, 2004). To avoid this problem, PROs should report regularly to

manufacturers and importers to discuss improvement of a material design and

alternative importing products that are more recyclable and/or that have less

environmental impacts. A brief potential structure of cooperation of a PRO and the

municipality in Mongolia is shown in Figure 9.

Fig 9. PROs and municipal participation in the EPR program (Mongolia)

• Monitoring by a third party

The OECD guidance (2001) stated the EPR program has to be monitored and

reported by a third party to improve quality and efficiency. One of the key reasons

for the successful EPR policy in Canada is that it addresses needs, evaluates

performances, and tracks operations. In British Columbia, an annual audited report

including recovery rate, revenues, and cost is submitted by the third party (Driedger,

2002), which means there will be no outside pressure. In Mongolia, the General

Agency of Specialized Inspection (GASI) is in charge of overall inspections, risk

assessment of health, and occupational safety. Therefore, the third party could be the

GASI or a NPO. In Mongolia, the importance of having the third party in the EPR

system is to monitor PRO operations and treatment technologies, and to report to the

 

 

26  

customs office about more suitable imported products having less hazardous

substances. Based on the valuable feedback from the third party, the national

guideline for imported products under the EPR policy will be improved with the

needs for different stakeholders being clearly addressed. Monitoring both the sale of

products as well as HHW in treatment facilities is one of the effective ways of

determining the EPR policy achievement (Driedger, 2002).

• Awareness of the EPR policy

Increasing the number of recycled products and materials in the market is

essential in order to achieve the EPR program goals. The market-driven program is

only effective for highly valuable recovered material; however, there is less valuable

material that has a high environmental impact. Therefore, some level of government

intervention is needed for increasing the use of recycled materials. In China, the

Government Procurement Law was legislated and implemented in 2003 to promote

purchasing and using green products in all government agencies. According to this

law, the purchased products must meet Chinese environmental protection standards

(Qiao and Wang, 2007). Table 7 shows energy saving by using recycled materials

instead of raw materials (Nnorom and Osibanjo, 2008).

Table 7. Energy saving from recycled materials

Material Energy saving (%)

Aluminum 95

Copper 85

Plastic >80

Iron and steel 74

Lead 65

Paper 64

Zinc 60

One of the ways of improving consumer awareness in HHW is by producers

labeling their products based on hazardous components and recycled material

proportions. The labels on hazardous components in products should include

information about potential toxicity and requirements for specific treatment. This will

raise consumer awareness of hazardous waste collection, recycling, and treatment. For

 

 

27  

instance, under the EU Battery Directive, all batteries containing hazardous chemicals

are required to be labeled “with crossed-out wheelie bin and chemical symbols”

(Mayers, 2007). Figure 10 shows the labels (Energizer, 2009). Moreover, this label on

products can be a tool for promotion of green purchases. The labels and

advertisements are producer`s responsibilities and the campaigns could be funded by

PROs.

Fig10. EU Battery Directive (labels)

Another way to increase consumer awareness of green products is

environmental education in public schools and advertising of green purchases through

television, radio, and newspaper. In Mongolia, the national broadcasting television

and public radio stations funded by the government should campaign for new

legislated policies and regulations, and against the environmental pollution caused by

HHW and effects on human health. Moreover, if producers and importers comply

with the EPR policy, the Mongolian National Broadcaster could advertise their

products and services for free through its environmental programs, could promote

companies` participation in the EPR program, and increase companies` social

responsibility. In Japan, children in kindergarten learn how to sort PET into the proper

trashcans. According to the elementary school curriculum in Mongolia (Minister

enactment, A/311, the Ministry of Education and Science of Mongolia, 2013), a

program about human interaction with the environment that extents until sixth grade.

The program includes natural phenomenon, the earth and the solar systems, plant and

fruit growths, forest, and atmospheric movement, etc. The program does not cover

practical activities to increase environmental awareness; however, practical programs

in the education system are necessary in both public and private elementary schools.

 

 

28  

Chapter 3

MERCURY RISK ASSESSMENT

Chapter 3 will identify current disposal methods, typically incineration, for

fluorescent bulbs containing elemental mercury in UB city. Given these disposal

technologies and the estimated releases mercury from the disposal into the

atmosphere, the risk of potential mercury exposure from municipal waste incineration

will be analyzed. The human health risk characterization part will focus on only

neurological risks for children by airborne exposure of elemental mercury vapor

(inhalation).

Mercury exposure from MSW incineration   Mercury exposure from MSW incineration is from batteries such as Hg-Cd

and Hg-Zn batteries, electric lighting such as fluorescent lamps and high intensity

discharge lamps, and dental amalgam (Velzen, Langenkamp, and Herb, 2002). In

1990 in the U.S., 70 percent of all produced mercury could be found in the waste

stream, which means 600 – 800 tons of mercury was disposed per year. At that time,

the total amount of MSW in the U.S. was around 200 million tons per year with 3-4

mg of mercury per kg waste (Velzen et al., 2002). According to Velzen et al (2002),

mercury in MSW treatment in the U.S. between 1991-1992 is illustrated in Table 8.

Table 8. Mercury sources in MSW

Batteries 67.3%

Fluorescent lamps 9.0%

Thermostats 8.7%

Plastics 5.4%

Latex paint residues 4.8%

Miscellaneous 2.9%

Food, paper 1.9%

In 1996, the Mercury-Containing and Rechargeable Battery Management Act

was enacted, which banned mercuric oxide batteries in the U.S.; since then, the

 

 

29  

mercury component in the waste stream (7.12 ton in 2010) has been decreasing and

non-mercury alternative batteries like alkaline manganese oxide batteries, alkaline

batteries, lithium ion button cell batteries and silver oxide batteries have increased in

market share (IMERC, Mercury use in Batteries, 2014). Li et al (2010) mentioned that

in China, mercury batteries have been decreasing in use since 2001; however, 153

tons of mercury was used in batteries in 2004.

Unlike mercury-containing batteries, the usage of fluorescent bulbs has been

increasing globally in recent years because of their energy efficiency and reduced

greenhouse gas emissions comparing with traditional incandescent lamps. In the U.S.

and EU, there is a ban on incandescent on (Lim, Kang, Oguseitan, and Schoenung,

2013). The similar growing pattern of fluorescent lamp (FL) use can be seen in

Ulaanbaatar city due to government promotion of FL usage in offices and public

places. However, the energy-efficient compact FLs have a potential mercury exposure

hazard. The mechanism of lighting up compact fluorescent lamps (CFL) can be

described as “the alternating current (AC) is used to make electrons collide with

mercury atoms to release ultraviolet (UV) light, which is changed to visible light

through the phosphor coating on the inside of the glass tube” (Lim et al., 2013). A

general structure of compact FLs is illustrated in following Figure 11 (Lim et al.,

2013).

Fig 11. Compact FL driver

AC: Alternating current, EMI: Electromagnetic interface, IC: Integrated current

 

 

30  

There have been the different amount of mercury and its amalgams in CFLs;

for instance, Raposo and Roeser (2001) mentioned that “the average amount of

mercury per lamp is being reported as 20.62 mg of Hg/lamp” in Brazilian fluorescent

lamp manufacturers. Based on their estimation, 1000 kg of Hg per year is released

into the environment in Brazil due to mercury-containing used lamps (Raposo and

Roeser, 2001). In the United States, “…according to the Association of Lighting and

Mercury Recyclers, approximately 670 million fluorescent bulbs were disposed of or

recycled in the United States in 2003.” (EPA, Fluorescent Lamp Recycling, 2009).

Due to an increase in usage of fluorescent lamps, in 2009 annually two to four tons of

mercury from FLs is released into the atmosphere in the U.S. (EPA, Fluorescent

Lamp Recycling, 2009). In China, approximately 200 tons of mercury is used in

fluorescent lamps and thermometers in each year. Over 90 percent of mercury

containing products end up in the MSW stream for treatment (Li et al., 2010).

The amount of mercury in a lamp is based on lamp types (Linear or CFL), a

brand, and wattage. Table 9 shows the amount of mercury in different kind of

fluorescent lamps (Nance, Patterson, Willis, Foronda, and Dourson, 2012).

Table 9. Mercury containing lamps

Country or region Lamp type Mercury (mg/lamp)

Europe Halophosphate lamps 10

Europe Triphosphate lamps 5-8

Canada Linear fluorescent tubes 3-50

USA Linear fluorescent tubes 1.4-50 (a) 0-100 (b)

Australia CFL 0.1-13

Canada CFL 1-25

United Kingdom CFL <10

USA CFL Avg. 4

USA CFL 5-50 (b)

a: Culver, 2008, b: NEWMOA

 

 

31  

Globally, the main forms of total mercury emission to the atmosphere are

gaseous elemental mercury (53%) and divalent mercury (37%). Carpi (1996) states

that, in general, 20-50 percent of elemental mercury (Hg0) and 50-80 percent of

divalent mercury (Hg II) result from coal combustion; and 10-20 percent of Hg0 and

75-85 percent of Hg (II) result from municipal waste incineration. For instance,

according to the EPA Emission Test Report (EPA/600/SR-93/181), the concentration

of mercury in Camden county MSW stack is 100-1000 µg/m3, all of which is divalent

mercury. With the exhaust steam, elemental mercury reacts with HCl and Cl2 or is

oxidized into HgO. Increasing the temperature and the concentration of HCl and Cl2

results in higher concentrations of oxidized mercury compounds and mercuric

chloride (HgCl) (Carpi, 1996). HCl and Cl2 are produced from chlorine-containing

plastic products in waste stream.

Studies in China showed that MSW incineration resulted in 94.39 percent of

mercury being in the fly ash (Zhou et al., 2015) and around 4 percent of the total

mercury is in the bottom ash (Carpi, 1996). According to Zhou et al (2015), the

average concentration of total mercury in fly ash from 15 different samples is 10

mg/kg; however, the total mercury in 7 samples from eastern China had a much

higher than the average concentration (10 mg/kg). Zhou et al (2015) mentioned this is

because rapid economic growth in that area led to an increase in mercury components

in MSW. Table 9 shows different forms of mercury emissions from waste

incineration in North America and Asia. As seen in Table 10, in both regions, divalent

mercury emission in gas state is dominant (E. G. Pacyna and J. M. Pacyna, 2002).

Table 10. Mercury emissions from waste incineration

North America Asia

Gaseous elemental mercury (Hg0)

13.3 ton 6.5 ton

Gaseous divalent mercury (Hg (II))

39.5 ton 19.6 ton

Particulate mercury 13.3 ton 6.5 ton

 

 

32  

In UB city, FLs from offices and consumer usages are disposed into MSW

stream. Based on different studies (Carpi, 1996, Zhou et al. 2015, E. G. Pacyna and J.

M. Pacyna, 2002), gaseous elemental and divalent mercury in MSW fly ash is a major

way for mercury to be released into the environment. Therefore this exposure is the

main pathway in UB city; moreover, the city`s solid waste disposal methods are open

burning or landfilling after incomplete incineration. Aucott et al (2003) estimated a

typical broken fluorescent bulb containing approximately 20 mg of mercury releases

3-8 mg of mercury in two weeks. Moreover, they found during FLs final disposal and

recycling processes approximately 1-80 percent of the total mercury is released from

the bulbs (Aucott et al., 2003). Another study (Johnson et al., 2008) reported that the

13W FL released 30 percent (1.34mg) of the total amount of mercury in the bulb, and

old lamps tend to release less mercury than a new one (Nance et al., 2012).

Figure 12 shows the map of the former municipal landfill in UB city, which is

close to residents and the location is 47°56'42"N 106°56'54"E (wikimapia). On the

map, every white dot represents Mongolian traditional house that has a stove inside

and it burns coal to heat the house in cold winter (average temperature in winter from

-33.4C to -18.9C) (WeatherSpark, Average Weather for Ulaanbaatar, Mongolia).

Fig 12. The former MSW facility

 

 

33  

Figure 13 shows how close the former MSW facility is to the city (center of UB city

coordination:  47°55'13"N 106°55'2"E). The facility was located in the northeast part

of the city and up in the mountains. In the city, the annual average wind direction is

from the northern side (23%) and the northwest side (16%) for a total of 39 percent,

which means 61 percent of the annual wind comes from the northern side

(WeatherSpark, Average Weather for Ulaanbaatar, Mongolia).

Fig 13. The MSW facility location and the city center

This landfill was overwhelmed with the waste load and was closed due to the amount

of waste from consumers and industries and a lack of land for disposal. Even though

it is closed currently, the main treatment for the waste is open burning and then

disposal in a landfill. Based on the average amount of 20 mg mercury (Raposo and

Roeser, 2001, Aucott et al., 2003) in a fluorescent lamp, the total amount of mercury

in imported FLs into the country can be estimated. However, it is not possible to

determine the amount of FLs in the city because the customs office manages and

reports only the total amount of imported lamps nationwide. Therefore, the amount of

FLs in UB city is estimated based on population, where at least half of the populace

lives in the city. Because of urbanization and population in the capital city, this risk

assessment assumes approximately 60 percent of the total FLs are used in UB city.

Aucott, McLinden, and Winka (2003) showed that when the temperature is between

5-30°C, approximately 17-40% of the mercury is volatilized from a broken

fluorescent lamp during a two-week period; therefore, open burning of FLs in the

 

 

34  

MSW facility will accelerate gaseous mercury emission into atmosphere. Table 11

shows the total amount of FLs and of mercury released from FLs in the MSW

treatment.

Table 11. The amount of mercury released into the atmosphere

Average imported FLs Mercury in the imported FLs

Mercury released into air

National 6447874×103 yr 20mgHg×6447874×103

=128957480×103 mg/yr 90%×128957.48kg =116,061.73 kg/yr

Ulaanbaatar city

6447874×103×60%=3868724×103yr

20mgHg×3868724×103

=77374488×103 mg/yr 90%×77374.488kg= 69,637 kg/yr

The elemental mercury vapor and divalent mercury exposure pathways are shown in

Figure 14. As seen in the figure, the most of divalent mercury vapor is deposited in

regional and local areas; therefore, residents who live near to the MSW incineration

are at greater risk of developing detrimental health effects by inhalation of mercury

vapors.

Fig 14. The metallic mercury airborne exposure pathway

 

 

35  

Hazard Identification

Physical and chemical properties of elemental mercury

Mercury is a transition metal, so it has different valence electrons, which

result in various complex compounds such as mercuric (II) chloride (HgCl2) mercury

(I) chloride (Hg2Cl2), and mercuric (II) acetate (HgC4H6O4) (Agency for Toxic

Substances and Disease Registry (ASTDR), 1999). Since elemental and inorganic

(metallic) mercury are the main forms inside FLs, this chapter and risk calculation

will focus on only their characteristics. Elemental and divalent mercury`s physical

and chemical properties are in Table 12 (ASTDR, 1999).

Table 12. Physical and chemical properties of mercury

Characteristic Mercury Mercuric (II)

chloride

Mercury (I)

chloride

Chemical formula Hg HgCl2 Hg2Cl2

EPA hazardous

waste

U151; D009 D009 No data

Molecular weight 200.59 271.52 472.09

Color Silver white (liquid

metal) tin white

(solid)

White White

Physical state Heavy, liquid metal Crystal, granule or

powder

Heavy powder,

rhombic crystal

Melting point -38.87°C 277°C 302C

Boiling point 356.72°C 302°C 384C

Odor Odorless Odorless Odorless

Solubility in water 0.28µmoles/L at

25°C

1 g/35ml,

6.9g/100cc at 20°C

2.0×10-4g/100ml at

25°C

Solubility in

organic solvents

2.7mg/pentane,

soluble in lipids

1g/3.8ml alcohol,

1g/200ml C6H6,

1g/22ml ether

1g/12ml glycerol

Insoluble in

alcohol, ether

LogKow 5.95 No data No data

 

 

36  

Vapor pressure 2×10-3mm Hg at

25°C

1mm Hg at 136°C No data

Henry`s law

constant at 24.8°C

No data No data No data

Degradation

reaction rate

constant

Gas-phase reaction

with ozone 1.7×10-

18 cm3/mol/s

No data No data

As shown in Table 12, elemental mercury (Hg0) has a high vapor pressure at

room temperature, which means that, during burning processes in a MSW plant, it is

easily vaporized to the atmosphere and becoming highly toxic gas. Hg0 is soluble in

lipids so it is stored in fatty tissues in humans and animals. Due to low water

solubility, the residence time Hg0 in atmosphere is 0.5-2.0 years (Carpi, 1996). In

general divalent mercuric salts are soluble in water; for instance, waster solubility of

mercuric (II) chloride is 1g/35ml H2O (ASTDR, 1999, and Bose-O`Reilly et al.,

2010).

Toxicology: Neurological effects in humans According to the Toxicological Profile for Mercury (ASTDR, 1999),

inhalation exposure is the primary route of concern for metallic mercury vapor

(elemental mercury) “70%-85% of inhaled mercury vapor is absorbed by the lungs

into the bloodstream” (Bose-O`Reilly et al., 2010). The main organs for the elemental

mercury exposure are the lung, kidneys, and the central nervous system.

Several studies have mentioned that a high level of mercury vapor can lead to death in

humans, which means the acute exposure (14 days or less) of mercury is highly toxic

(ASTDR, 1999). “Minimal Risk Levels (MRLs) are derived for acute (1–14 days),

intermediate (15–364 days), and chronic (365 days and longer) durations and for the

oral and inhalation routes of exposure.” Table 16 shows Lowest-Observed-Adverse-

Effect Levels (LOAELs) and No-Observed-Adverse-Effect Levels (NOAELs) for

chronic exposure to mercury for some animals and humans (ASTDR, 1999).

 

 

37  

Table 13. NOAEL and LOAEL of animals and humans

Species Exposure duration System NOAEL

Rat 71wk

5d/wk

7hr/d

Renal 0.1 mg/m3

Dog 83wk

5d/wk

7hr/d

Renal 0.1 mg/m3

Rabbit 83 wk

5d/wk

7hr/d

Renal 0.1 mg/m3

Species Exposure duration System LOAEL

Human 1-5yr Neurological 0.076 mg/m3

(Male)

Human 1-41yr (mean:

15.3yr)

Neurological 0.026 mg/m3

(Male)

Human 0.7-24yr Neurological 0.014 mg/m3

According to the toxicology profile for mercury (ASTDR, 1999),“The central nervous

system is probably the most sensitive target organ for metallic mercury vapor

exposure”. McFarland and Reigel (1978) showed that exposure (less than 8 hours) to

elemental mercury vapor concentration at 44 mg/m3 gave workers “long-lasting

feelings of irritability, lack of ambition, and lack of sexual desire” (Toxicology profile

for mercury, 1999).

Based on Table 13 data, Figure 15 illustrates neurological and renal effects on both

humans and animals due to chronic exposure of elemental mercury vapor (ASTDR,

1999).

 

 

38  

Fig 15. Dose – renal and neurological effects of mercury airborne exposure

At high concentrations of mercury vapor exposure over 2 days, a “54-year-old man

exhibited a syndrome resembling amyotrophic lateral sclerosis, characterized by

slowed conduction velocities (suggestive of peripheral nerve damage)” (Toxicology

profile for mercury, 1991). According to Rowens (1991), low-level chronic exposure

to elemental mercury affects the neurologic system resulting in “symptoms of tremor,

neuropathy, and changes in personality referred to as mercurial erethism” (ASTDR,

1991).

The toxicology profile for mercury (ASTDR, 1991) states “…chronic-duration

exposures to metallic mercury vapor have resulted in tremors (which may be mild or

severe depending on the degree of exposure), unsteady walking, irritability, poor

concentration, short-term memory deficits, tremulous speech, blurred vision,

performance decrements in psychomotor skills (e.g., finger tapping, reduced hand-eye

coordination), paresthesias, decreased nerve conduction, and other signs of

neurotoxicity.” Figure 15 shows the mercury vapor concentration in the air and its

human neurological effects (Friberg, WHO, 1991).

 

 

39  

Fig 16. Dose-response of elemental mercury vapor exposure (inhalation)

Mercury vapor concentration (mg/m3) 1: Control group, 2: <0.01-0.05, 3: 0.06-0.1, 4:

0.11-0.14, 5: 0.24-0.27

Metallic mercury vapor can damage on the kidneys and respiratory   systems   in  

humans.  The  following  shows the effects on these systems.

1. Renal  effects  

  The proximal tubules are the major target in the kidneys. A study shows (NTP,

1993) mice were given 20 mg Hg/kg/day in five days in a week increased “kidney

weights at 3.7 mg Hg/kg/day and acute renal necrosis at 59 mg Hg/kg/day” (ASTDR,

1999). Due to a high dose exposure of elemental mercury, the tubular cells are not

able to regenerate, meaning the function of the kidneys is going to fail (ASTDR,

1999). The study (Nielsen et al. 1991) on mice shows 10mg/Hg/kg of mercuric

chloride led damages to renal tubular cells (ASTDR, 1999).

2. Respiratory  Effects  

Several studies have reported death in humans following acute exposure to

high concentrations of elemental mercury vapor (Campbell 1948; Kanluen and

Gottlieb 1991; Matthes et al. 1958; Rowens et al. 1991; Soni et al. 1992; Taueg et al.

1992; Teng and Brennan 1959; Tennant et al. 1961). The deaths were correlated to

respiratory failure due to high-level exposure of mercury. The common symptoms are

“cough, dyspnea, and tightness or burning pains in the chest” (ASTDR, 1999).

According to the Mercury toxicology profile (1999), the lungs exposed to mercury

vapor show “diffuse infiltrates or pneumonitis” and “impaired pulmonary function.”

0  

5  

10  

15  

20  

25  

30  

35  

1   2   3   4   5  

Sym

ptom

s (%

)

Nervous disorders

 

 

40  

Risk Characterization

Elemental mercury risk calculation  

1. Mercury concentration in the air

The total amount of mercury in the air (Table 10) is estimated to be 69,637kg

Hg/yr (190,786.30g Hg/d). To estimate the mercury concentration in air, this

assessment used the mean yearly mixing height over the urban area since municipal

solid waste is burned outside in the city. Therefore, we need to have the total volume

of air around the former MSW. Chen et al (2012) used the zone of 0-500m distance

from the pollutant source (the lead-acid battery recycling factory) as a pollutant area;

therefore, the zone can be used in our risk calculation is approximately within the 500

m radius. The actual 500 m radius from the former MSW plant is illustrated in Figure

17.

Fig 17. The pollutant zone within the 500 m radius

Several studies have used average mixing heights to calculate pollutant concentrations

in urban areas, as shown in Table 14 (Holzworht, U.S. EPA, 1972; Benarie, 1974).

UB city`s has an annual average wind velocity of 8.1 mile per hour (3.62 meter per

second) (WheatherSpark), which is close to the mean wind speed in Tokyo (3.7 m/s)

 

 

41  

so this assessment used Tokyo as the example city for the assumptions of a mixing

height of 830m (±100m) (Benarie, 1974).

Table 14. The mean wind velocity and average mixing height

Cities Mean wind velocity

(meter per second)

Average mixing height

Los Angeles 2.14 m/s * 680 m

New York 5.45 m/s * 1060-1080 m

Las Vegas 4.06 m/s * 1382 m

Pittsburg 3.97 m/s * 1010 m

Washington D.C. 4.2 m/s * 1049 m

Paris 4.1 m/s 690 m

Bordeaux 3.1 m/s 745 m

Tokyo 3.7 m/s 830 m

Osaka 2.9 m/s 920 m

Lyon 3.3 m/s 620 m

Ulaanbaatar 3.62 m/s 830m (±100m) * Wind-spees.weatherdb.com

Based on the mixing height (Table 20) and the radius of the polluted zone (Figure 17),

the total volume of the contaminated area is:

V= πr2h = 65155×104 m3

Therefore, the mercury concentration per day in the contaminated area is:

Hg concentration = MassVolume

= 190,786.30g Hg65,155×104 m3

= 0.2928mg Hg/m3

2. Chronic exposure of neurological damages in humans (Table 2)

Exposure duration: 0.7-24yr

LOAEL concentration: 0.014 mg/m3

3. Reference concentration (RfC)

RfC = LOAELUF

UF (Uncertainty Factor): 100

RfC = 0.014 mg/m3

100 = 0.00014 mg/m3

 

 

42  

4. Chronic daily intake (CDI) by inhalation

CDI = CA×IR×ET×EF×ED

BW×AT

CA: Contaminant concentration in air: 0.2928mg Hg/m3

IR: Inhalation rate: 0.019m3/hr (children)

ET: Exposure time: 12hr/day (only daytime)

EF: Exposure frequency: 365d/yr (MSW treatment operation time)

ED: Exposure duration: 70yr (lifetime exposure)

BW: Body weight: 33kg (6-12yr children)

AT: Averaging time: 365d/yr×70yr

CDI1= 0.2928mg Hg/m3×0.019m3/hr×12hr/d×365d/yr×70yr

33kg×365d/yr×70yr

CDI1=0.00202 mg/kg/d

When exposure time is 24hr per a day, chronic daily intake is:

CDI2=0.00404 mg/kg/d

5. Hazard Quotient (HQ)

HQ= CDIRfC

HQ1= 0.00202 mg/kg/d0.00014 mg/m3

= 14.42

HQ2= 0.00404 mg/kg/d0.00014 mg/m3

= 28.85

The both hazard quotients are bigger than one meaning there is a potential hazard for

children`s neurological damages.

 

 

43  

Chapter 4

LEAD RISK ASSESSMENT

Chapter 4 will analyze and summarize current recycling methods and

technologies for used lead-acid vehicle batteries containing lead in UB city. The last

part of the chapter will cover the lead hazard identification and risk calculation based

on the airborne exposure for lead from used LABs recycling factories. Even though

the main health concerns of lead toxicity is the nervous system, to identify health

effects for different organisms the renal effects on children will be discussed in

toxicology of the hazard identification.

Lead exposure from lead-acid battery recycling factors Lead-acid batteries are used for various, such as powering/starting heavy

trucks, buses, boats, and backup power for building lightings and electronic

equipment. In 2015, the total lead production in the U.S., Canada, and Mexico is

estimated to be 2,090,000 tons. In this estimation, recycled lead consumption is

approximately 86 percent (1,800,000 tons) of all lead consumption (Queneau, Leiby

and Robinson, 2015). Moreover, 100 million lead acid car batteries were sold in the

U.S. and Canada in 1999 and 1,000,000 tons of lead were used for these automobile

batteries in the U.S. at the same time. Total usage of recycled lead in the U.S. was 76

percent of all produced lead in 1999 (Commission for Environmental Cooperation,

2007). According to the estimation, by 2015, lead from used lead-acid batteries will

account for 95 percent of total recycled lead in North America (Queneau et al.

2015). Battery industries use more than 75 percent of all lead consumption globally.

Table 15 shows the consumption of lead for batteries in different countries (Basel

Convention and UNEP report, 2003).

Table 15. Consumption of lead for batteries (1993)

Country Percentage

U.S. 83

Japan 69

France 65

 

 

44  

Germany 56

Italy 46

UK 34

In China, there are approximately three hundred lead recycling plants in 2013

and most of them are located in Jiangsu, Shandong, Henan, Hubei, Hunan, Hebei, and

Anhui provinces. These recycling factories produce 80 percent of all lead production

in China. The annual lead consumption for lead-acid batteries from secondary lead

was approximately 60 tons in 2000 and it has been increased by approximately 100

tons in 2010 (Zhang, 2013).

There are different types of batteries based on their purpose; however, as

shown in Figure 18, a typical lead-acid battery has positive and negative terminals,

sulfuric acid solution, positive and negative plates, and plate separators (Basel

Convention and UNEP, 2003).

Fig 18. Internal structure of a lead-acid battery

“(1) and (2): terminals made of lead; (3): one for each battery element, where

distillated/deionized water can be replaced; (4): made of lead that makes electrical

contact between plates of same polarity; (5) and (11): originally made of ebonite, but

now more commonly made from either polypropylene or co-polymer; (6): the

electrolyte of the battery; (7): usually a part of the box, provide chemical and

electrical isolation between the electrical elements; (8): made of PVC or other porous

 

 

45  

materials, avoid physical contact between two contiguous plates but, at the same time,

allowing free movement of ions in the electrolyte solution; (9): constituted by a

metallic lead grid covered by a lead dioxide (PbO2) paste; (10): constituted by

metallic lead plates; (11): a series of negative and positive plates;” (Basel Convention

and UNEP report, 2003).

In the positive terminals, lead dioxide (PbO2) is converted to lead sulfate (PbSO4) and

in the negative terminals, elemental lead is converted into lead sulfate (PbSO4).

Sulfuric acid (H2SO4) acts as the electrolyte, providing the SO4- ion to both reactions.

The chemical reaction in a lead battery is seen in Figure 19 (Basel Convention and

UNEP report, 2003 and Vest, 2002).

Fig 19. The chemical reaction in a lead-acid battery

After the discharge-recharge process, the lead dioxide (PbO2) plates become

contaminated by lead sulfate (PbSO4), resulting in the eventual loss of the ability to

operate. . In the bottom of the battery, “a sludge layer” containing 55-60 percent of

PbSO4, 20-25 percent of PbO2, 1-5 percent of PbO, and 1-5 percent of metallic Pb

develops (Basel Convention and UNEP report, 2003).

The lead-containing component varies based on the brand, battery size, and

usage or purpose. Generally the amount of lead is approximately 60 percent to 80

percent of the weight of a lead-acid battery. “An average automobile battery weights

17.2 kg, and contains 8.6-9.1kg of lead” (Office of Emergency and Remedial

Response, EPA, 1992). Lead percentages in varies types of batteries are shown in

Table 16 (Basel Convention and UNEP report, 2003 and Vest 2002).

Table 16. Lead components in automobile batteries

Battery type Lead component Total weight

12V-44Ah-210A-starter battery in a

hard rubber casing

58.8% (8.82 kg) 15kg

12V-44Ah-210A-starter battery in a 63.9% (8.62kg) 13-14 kg

 

 

46  

PP casing

12V-44Ah-220A vehicular battery 63.2% (8.4kg) 13.3 kg

As seen in table 13, the battery`s lifetime varies in different countries and

regions; typically in developing countries, it averages 20-24 months. The useful life

of batteries depends on different factors such as: “incomplete charging process;

remaining too long without use, hot weather, deep discharging process, and low

electrolyte level “ (Basel Convention and UNEP report, 2003). As seen in Table 17,

India estimates that its battery lifetime is 1.8 years, which may be a result of the

quality of the batteries and the extreme hot weather.

Table17. Lifetime of automobile batteries (1995)

Country/Region Lifetime (yr)

Western Europe 5.3

Canada 5.0

Japan 4.5

Australia 3.1

USA 3.0

Brazil 2.4

India 1.8

Based on the above estimation (Table 12), 60 percent (8kg) in a weight of an

automobile battery in Mongolia can be considered as lead. Generally, Mongolia

imports most of its automobiles from Japan due to their high fuel efficiency. UB city

covers 0.3 percent (4704.4 km2) of all the land of Mongolia; however, due to the

rapid urbanization, in 2014, 60.9 percent (411,408) of all vehicles including buses,

trucks, and automobiles are in the city (Ulaanbaatar city Traffic Police, 2014).

According to Table 13, the lifetime of lead-acid batteries in Mongolia can be

estimated to be 2 years; therefore, a half of the total vehicles in the city will dispose of

used lead-acid batteries (LABs). Based on the above estimation, the total discharge of

lead from used LABs per year in UB city is shown in Table 18.

 

 

47  

Table 18. The annual amount of lead from LABs (Ulaanbaatar city)

Total automobile

batteries in UB city

(2014)

Total amount Pb in

automobile batteries in UB

city

Discharge of Pb based on

battery lifetime

411,408 411,408×8kg=3,291,264kg 3291264×0.5=1,645,632 kg/yr

According to Table 18, approximately 1645 tons of lead only from vehicle

batteries is going to be disposed in every year in the city. In Mongolia, several small-

scale recycling lead-acid battery factories exist throughout the country; however, the

specific data and information about the recycling process is not publicly available due

to the company`s secret processes. The lack of information means that the treatment

or lack of treatment cannot be determined. According to our interviews (2015) of

automobile importers and small repair businesses in UB city, most of the recycling

factories are in the countryside where the inspection and environmental regulations

are less restrictive than in the city. Therefore, we assume that the recycling process in

Mongolia is similar to a general process of used LABs recycling in developing

countries. Figure 20 shows the recycling stages of a lead-acid battery in general

(Vest, 2002).

Fig 20. Typical recycling process of used LABs in developing countries

As shown in Figure 20, after dismantling the batteries, the grids and lead pastes are

separated in the screening stage and processed individually because grids are in metal

 

 

48  

stages. Therefore, the metallurgical process for converting grids (PbO/PbSO4) to pure

lead (Pb) can be shortened and it is melted at 500°C (Vest, 2002). Gas emission

control devices are not installed in most of small factories due to financial constraints

and less environmental regulation and monitoring in Mongolia; therefore, potential

hazardous gas emissions and particulates with Pb, Sb, As, and SO2 could be emitted to

the atmosphere during the smelting and refining processes (Vest, 2002 and Basel

Convention and UNEP report, 2003). Lead contaminant dusts from furnaces can

contain up to 65 percent of lead. During the manual battery breaking process, acid

electrolyte and lead contaminated dust can be emitted directly to the environment.

Moreover, an average approximately 300-350kg of slag per ton of metallic lead can

be produced during the lead reduction process (melting process) and around 5 percent

(w/w) of the slag is lead compound (Basel Convention and UNEP report, 2003).

Due to a lack of technologies and expertise, lead exposure from used LABs

recycling factories can be more common in developing countries. For instance,

according to a study (Chen et al., 2012), “dust lead concentration within the factory

area varied from 441,460 to 708,560 mg/kg with a mean value of 601,030 mg/kg.”

The recycling lead-acid battery factory in the study has a 1 million kilovolt-ampere-

hour (kVAh) battery production capacity. Chen et al (2012) states lead concentration

in outdoor dust samples collected from polluted areas (0-500m distance from the

factory) was from 270 mg/kg to 19,490 mg/kg (mean: 9,250 mg/kg). Another study

(Were et al., 2012) shows airborne lead levels in breaking and the furnace

(temperature 1000-1200°C) of a recycling factory were 420±127 µg/m3 and 447±154

µg/m3, respectively. The concentration of lead in the air exceeded Occupational

Safety and Health Administration Permissible Exposure Limit for lead, which is 50

µg/m3 (8-hour workday). Chen et al (2012) show that the lead concentration in the

dust decreased when sampling distance from the factory increased. In addition, the

study suggested the main lead pollutant source for the neighboring environment is

lead particulates in the atmosphere. Lead particulate in the lead powder in the factory

is 90 percent inhalable particles (<10µm), 50 percent fine particles (<2.5µm), and 25

percent ultrafine particles (<1.0µm) (Chen et al., 2012). After analyzing lead dust in

the cloth bag filter of the air treatment process, Chen et al., (2012) found that 32

 

 

49  

percent of inhalable particle was captured and approximately 60 percent of the lead

particulate was released to the surrounding environment.

However, in Mongolia, air control devices are not used in most family-owned

recycling factories; thus, the lead particulates in the factory air are going to be

released into the atmosphere directly. According to the Office of Emergency and

Remedial Response (OERR) (U.S. EPA, 1992), lead contaminant concentration in

used LABs sites is up to 7 percent of the total lead in a battery. Pb, PbS, PbSO4, and

PbOx are the main pollutants in the recycling factories (OERR report, 1992, and

Office of Research and Development, U.S. EPA, 2006). Baldasano et al., (1997)

shows that “Pb smelters had mean emission factors of 0.1 grams and 0.05 grams of Pb

emitted per kg of Pb processed for primary and secondary Pb smelters respectively”

(Office of Research and Development, U.S. EPA, 2006). Therefore, we can assume

every one kg of lead processing in the recycling factory in Mongolia will emit 0.05 g

of Pb to the environment.

The total amount of the lead released is:

1,645,632 kg Pb/yr × 0.05 g Pb emission/kg Pb=82,281.6 g of Pb/yr

For the worst-case scenario, 82.28 kg of lead per year can be emitted to the

atmosphere directly. Even though a part of the lead is in the slag, eventually lead dust

and lead particulates (inhalable particulates <10 µm) will be released to the air

because of not having waste treatment for the slag.

Hazard identification

Physical and chemical properties of lead

Lead is in the Group 14 (IVA) of the periodic table, which has electron

configuration of (Xe) 4f14 5d10 6s2 6p2. In nature, lead compounds are one of four

stable isotopes: “208Pb (51–53%), 206Pb (23.5– -27%), 207Pb (20.5–23%), and 204Pb

(1.35–1.5%)” (ASTDR, 2007). In the environment, lead exists commonly as Pb (II).

Pb (IV) exists rarely under normal environmental conditions. Emissions from

secondary smelters typically are Pb, PbSO4, PbS, and PbOx. Table 19 shows the

physical and chemicals properties of elemental lead, lead sulfate and lead sulfide

(Lead Toxicology Profile, ASTDR, 2007).

 

 

50  

Table 19. Physical and chemicals properties of lead

Characteristic Lead Lead sulfate Lead sulfide

Chemical structure Pb PbSO4 PbS

EPA hazardous waste D008 No data D008

Molecular weight 207.20 303.26 239.27

Color Bluish-gray White Black, blue or

silver

Physical state Solid Solid Cubic, metallic

crystal

Melting point 327.4C 1170C 1114C

Boiling point 1740C No data 1281C

Density at 20C 11.34g/cm3 6.2 g/cm3 7.57-7.59 g/cm3

Odor None No data No data

Solubility water at

25C

Insoluble 42.5 mg/L 0.86 mg/L at 13

C

Organic solvents Insoluble Insoluble alcohol Soluble in nitric

acid, insoluble in

alcohol

Vapor pressure 1.77 mmHg at

1000C

No data 10 mmHg at

975C

Henry`s law constant No data No data No data

 

 

51  

The average residence time of lead in the atmosphere is 10 days, which allows lead to

travel long distances through the air. Lead is insoluble in water; however, wet and

dry deposition is the main removal mechanism for lead particles from the atmosphere.

Therefore, the major contaminant source is soils. According to a study (Corrin and

Natusch 1977), “the median particle distribution for lead emissions from smelters is

1.5 µm with 86 percent of the particle sizes under 10 µm” (ASTDR, 2007). Lead

becomes water-soluble only at certain pH values. At pH higher than 5.4, lead

solubility in water is about “30 µg/L in hard water and approximately 500 µg/L in soft

water” (ASTDR, 2007). “Plants and animals may bioconcentrate lead”, however, soil

organic matter absorbs lead; therefore, bioavailability of lead is limited (ASTDR,

2007).

Toxicology: Renal effects in humans   According to the Toxicology Profile for Lead (TPL) (ASTDR, 2007), there is

no safe threshold for lead toxicity in humans; therefore, “MRLs were not derived for

lead”. An epidemiology study shows (Haefliger et al., 2009) the blood lead

concentration of 50 children near used LABs recycling factories in Senegal ranged

39.8-613.9 µg/dL (mean: 129.5 µg/dL) and 17 children out of 50 suffered from

neurological damage from lead toxicity. Heafliger et al (2009) found that 42 percent

of the total children had an increased incidence of severe gastrointestinal and 34

percent of the children had a neuropsychiatric disorder. The indoor lead concentration

in soils was 14,000 mg/kg and the outdoor concentration was 302,000 mg/kg. The 18

children (1-6 years old) died due to “encephalopathy resulting from severe lead

intoxication” (Heafliger et al., 2009). LOAELs for neurological disorders in adults

and children are shown Table 20 because the most sensitive organism of lead toxicity

is nerve systems (WHO, 2010).

Table 20. LOAELs for neurological disorders

LOAELs (µg/L) Effects

300 (adults) Peripheral nerve dysfunction

150-200 (females)

200-300 (males)

Erythrocyte protoporphyrin elevation

100-150 (children) Vitamin D3 reduction

 

 

52  

150-200 (children) Cognitive impairment, Erythrocyte protoporphyrin elevation

A study shows (Cooper, 1988) 4,519 battery plant workers who worked between 1946

and 1970 had greater mortality rates (p<0.01) compared to the national average.

During the period 1947 - 1972, the mean lead blood level was 63 µg/dL for 1,326

battery plant workers (ASTDR, 2007).

“Classic lead nephrotoxicity is characterized by proximal tubular nephropathy,

glomerular sclerosis, and interstitial fibrosis and related functional deficits, including

enzymuria, low- and high-molecular weight proteinuria, impaired transport of organic

anions and glucose, and depressed glomerular filtration rate” (TPL, ASTDR, 2007).

The overall dose-response curve suggests “an increasing severity of nephrotoxicity

associated with increasing” blood lead (PbB) concentration in humans. The studies

states severe lead poisoning in early childhood will cause renal failure or reduce the

kidney function (P. Castellino, Bologna and N. Castellino, chapter 12, 1995).

According to the toxicology profile for lead (ASTDR, 2007), a longitudinal study

shows a significant relationship between increasing serum creatinine concentration

and an increase in blood lead level (PbB <10 µg/dL). Furthermore, a cross-sectional

analysis resulted risks of chronic renal disease (“defined as severely depressed

glomerular filtration rate”) in association with blood lead level (PbB <6 µg/dL). Table

21 shows the renal effects at various PbB concentrations (TPL, ASTDR, 2007).

Table 21. Lead nephrotoxicity in humans (Dose-Response)

Response Dose in PbB

Effects on glomerular filtration

<10 µg/dL

Effects on enzymuria and proteinuria

>30 µg/dL

Severe deficits in function and pathological changes

>50 µg/dL

According to the TPL (ASTDR, 2007), the target-organ toxicity dose for renal in

humans is PbB concentration at 40-100 µg/dL (chronic nephropathy). For the child

this does, PbB level is approximately 34 µg/dL (Verberk et al., 1996).

 

 

53  

Risk Characterization

Lead risk calculation

1. Lead concentration in the air

The total amount of lead released to the atmosphere (p.23): 82,281.6g Pb/yr (225.43g

Pb/d). To get the lead concentration in the air, the same calculation is from the

mercury assessment is repeated for lead. Therefore, lead concentration per day is:

Lead concentration = 225.43g Pb/d65,155×104×m3

 = 0.000346mg Pb/m3

2. Chronic exposure of renal effects in humans

A study (Were et al., 2012) shows that an increase in airborne lead exposure resulted

in increasing PbB in employees who work in the used LABs recycling factory in

Kenya. Table 22 shows the airborne lead concentration inside recycling factory and

PbB concentration in workers (Were et al., 2012).

Table 22. Lead concentration in the air and PbB

Sampling location Airborne lead level (µg/m3)

Mean ± Standard Deviation

Blood lead level (µg/dL)

Mean ± Standard Deviation

Breaking 420±127 Not sampled

Furnace 447±154 61.6±14.7

Casting 430±101 63.9±11.7

Refining 405±137 61.1±9.1

All production 427±124 62.2±12.7

Office 59.2±22.7 43.4±6.6

Since there is no LOAEL, NOAEL for renal effects from the lead exposure by

inhalation, we will assume lead concentration from the air in the office of the used

LABs recycling factory 59.2±22.7 µg/m3 is as a LOAEL for nephrotoxicity in

humans. In the lead risk assessment, it considers renal effects in children who live in

500 m radius from the LABs recycling factories, thus the airborne lead level from the

office (Table 22) could be used for the reference concentration calculation. The blood

lead level 43.4±6.6 µg/dL from workers in the office is much higher than 10 µg/dL

 

 

54  

(Table 21), which indicates that lead concentration in the air (59.2±22.7 µg/m3) could

affect glomerular filtration in the human kidneys.

3. Reference concentration (RfC)

RfC = LOAELUF

LOAEL: 59.2±22.7 µg/m3 (Table 22), UF: 100

RfC = 59.2 µg/!!  

!"" = 0.592 µg/m3

4. Chronic Daily Intake (CDI) by inhalation

CDI = CA×IR×ET×EF×ED

BW×AT

CDI1=  0.000346mg Pb/m3×0.019m3×12hr/d×365d/yr×70yr

33kg×365d/yr×70yr

CDI1= 0.00000239 mg/kg/d

When ET is equal to 24hr/d, chronic daily intake is:

CDI2= 0.00000478 mg/kg/d

5. Hazard Quotient (HQ)

HQ= CDIRfC

HQ1= 0.40×10!! (<1)

HQ2= 0.82×10!! (<1)

The hazard quotients are less than one, which means there are likely to be no renal

effects on children who live near the used LABs factories; however, the factories do

not have waste treatment for the slag. Thus there could be environmental pollution to

groundwater contamination and the soil due to slag disposal.

 

 

55  

Chapter 5

CONCLUSIONS  

Implementing the EPR policy in a developing country has several benefits,

including reducing environmental and human health risks due to poorly designed

landfills and waste incinerators, and decreasing the municipal burden for the physical

and financial investment in waste management. Also the program improves the

producer responsibility for their products` waste, and promotes less virgin natural

resource use for a product and process. Manufacturers are the only important player

that can drive the design improvements to reduce hazardous substances from

products. Therefore, the overall costs from authorities for waste management can be

reduced under the EPR program. In addition, due to implementation of the EPR

program, the collection and recycling rates have been significantly increased and this

could be mimicked in UB city; for example, the recycling rate of packaging and

containers in Japan was increased by 27% from 1997 to 2000 (OECD, 2001).

As a developing country, the Mongolian government has an important role to

play in the development and implementation of HHW management and the EPR

policy nationwide. China has developed several regulations, environmental standards,

and promotion laws that are run by the government and that showed dramatic

improvement in the waste reduction. Under the EPR policy in Thailand, customers are

paid if they bring personal computers and notebooks to the certified retailers or a

collection center (Kojima et al., 2009). There should be the government intervention

in the waste management system if there will be a potential hazard for the residents

and the environment.

In 1999, Bennett at the OECD EPR workshop introduced the following matrix

system (Table 23) to give a brief guide to policy makers about waste management

under the EPR framework (OECD, 2001). The horizontal and vertical axes represent

environmental impacts and the cost of recovery value, respectively.

Table 23. Matrix application for the EPR policy framework

Matrix A

Market driven EPR program

Matrix B

Voluntary EPR program

 

 

56  

As shown in Matrix A, the high recovery rate for waste that can be voluntary

collected by individual producers can be explained because it is a market-driven

system and the recycled material is highly valuable. This system fits in used LABs

recycling process. The risk assessment for lead exposure from recycling factories was

low compared to the mercury hazard quotient. This may be because most of the lead

contaminant is in the slag or wastewater from the factories. In the used LABs

processing stage, the lead particulate matter in the air is higher than outside; therefore,

the occupational human health risk assessment would have been a bigger issue.

However, because lead is a valuable metal, factories are still continuing recycling

used LABs and putting the employees` health to danger. In this case, the

environmental impacts and human health effects cover a wide range; therefore, to

prevent these risks the mandatory EPR program operated by PROs for used LABs is

essential in Mongolia.

Based on the children’s health risk assessment for elemental mercury exposure

as it relates to potential neurological damages, authority participation in an EPR is

necessary because the calculated results in Chapter 3 showed a high risk to the

Positive value: post-consumer

phase

Positive value

Low High

Low environmental impact

waste

Negative

Value

Low High

Matrix C

Negotiated EPR program

Medium impact waste

Negative

value

Low High

Matrix D

Mandatory EPR program

High environmental

impact waste

Low High

Negative

value

 

 

57  

children’s neurological system from the current system of open burning of waste.

Unlike used LABs where airborne exposure to lead is minimal even from open

burning and where lead recovery is valuable, FLs do not have highly valuable

materials (mercury content is less than in e-waste), the mandatory EPR program is

needed to regulate FL waste. Therefore, this kind of HHW management fits in the

Matrix D, which shows that government intervention is important. The risk

assessment calculations can be used by the government to demonstrate the health

concerns to the citizens to encourage participation in the EPR program.

 

 

58  

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