Table of Contents
Page
1.0
1.1
1.2
2.0
2.1
2.2
3.0
3.1
3.2
3.3
3.3
3.4
4.0
4.1
4.2
4.3
INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Causality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
The Risk Function . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2.1
Reference level for air pollution: Impact above what level
of pollution? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2.2
What Are the Sources of Airborne Particles? . . . . . . . . . . . .
1.2.3
Health Effects of Particulate Matter . . . . . . . . . . . . . . . . . . .
1
1
1
2
4
5
WHAT IS PARTICULATE MATTER? . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
What Are the Physical and Chemical Characteristics of
Particles? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
2.1.1
Size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
2.1.2
Composition of Particulate Matter . . . . . . . . . . . . . . . . . . . . 9
Ambient Particulate Matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10
PM AND RESPIRATORY HEALTH RESPONSES . . . . . . . . . . . . . . . . 12
The Respiratory Tract as a Target . . . . . . . . . . . . . . . . . . . . . . . . . . 12
3.1.1
Exposure to Particles and Effects on Cells and Tissues . . 12
3.1.2
Attributes of Particles That Are Associated with
Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16
Establishing a Dose-Response Relationship for Effects of Particulate
Matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16
How Do Particles Cause Effects? . . . . . . . . . . . . . . . . . . . . . . . . . . 18
3.3.1
Inflammatory Responses in the Airways . . . . . . . . . . . . . . . 18
3.3.2
Systemic Inflammatory and Other Vascular Responses. . . 19
Neural Control of Heart Function . . . . . . . . . . . . . . . . . . . . . . . . . . 20
Mechanisms of Particle Effects: Conclusions . . . . . . . . . . . . . . . . . 21
3.4.1
Are Some Individuals or Groups Particularly Susceptible to
the Effects of Particles? . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22
LINKAGE OF HEALTH IMPACTS WITH ENVIRONMENTAL
POLLUTANTS THROUGH EPIDEMIOLOGIC STUDIES. . . . . . . . . 23
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23
Hospital Admission Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24
4.2.1
Assessment of Deaths Attributable to Air Pollution: LongTerm Predictions from Time-Series Studies . . . . . . . . . . . 25
4.2.2
The concept of death in time-series and cohort studies . . . 25
4.2.3
Four categories of air pollution-attributable deaths . . . . . . 26
4.2.4
Air pollution attributable death in time-series studies. . . . 27
Air pollution attributable death in cohort studies . . . . . . . . 28
4.3.1
Attributable number of deaths . . . . . . . . . . . . . . . . . . . . . . 28
RESOURCE DOCUMENT: The health assessment of particulate matter with special emphasis on mobile sources of
emission.
February, 2003 • CANTOX ENVIRONMENTAL
4.4
4.4.1
4.4.2
4.4.3
5.0
5.1
5.2
6.0
6.1
6.2
7.0
Mortality and Hospitalization: . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Recent Epidemiology Showing an Association of Mortality
with Urban Air Quality. . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Morbidity, Hospitalization and Urban Air Quality . . . . . . .
Urban Air and Increased Frequency of Admissions for
Cardiovascular Disease . . . . . . . . . . . . . . . . . . . . . . . . . . . .
CHRONIC EXPOSURE TO PM10 AND RESPIRATORY HEALTH . .
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Health Impacts from Long-term Exposure to Air Pollution . . . . .
5.2.1
Respirable Fine Particulate Matter PM2.5 . . . . . . . . . . . . . .
5.2.2
Inhalable Particulate Matter PM10 and Criteria
Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.2.3
What endpoints describe respiratory health and chronic
exposure of children and adults to particulate matter? . . .
5.2.4
What are the associations between other clinical effects in
children and chronic exposure to urban air based on PM?
5.2.5
What Are the Effects of PM10 on the Lung Function and
Incidence of Respiratory Symptoms in Adults? . . . . . . . . .
5.2.6
Sensitive Populations for Respiratory Impacts . . . . . . . . . .
28
29
31
32
34
34
36
36
37
39
48
49
50
SOURCES OF PM AND ASSOCIATED HEALTH EFFECTS . . . . . . . 51
Are all sources of PM10 equally responsible for health effects? . . . 51
What is the relationship between traffic and evidence for adverse
health effects from PM? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52
7.3
EXPOSURE ASSESSMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55
Personal Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55
What are the relationships between outdoor ambient monitors or
predicted levels of PM or VOC and actual human exposure?. . . . .56
Personal Exposure Monitoring and PM . . . . . . . . . . . . . . . . . . . . . 57
8.1
8.2
8.3
INTERACTION AMONG POLLUTANTS AND MEASUREMENT OF
EFFECT ON HEALTH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60
Interactions of Pollutants and Allergens . . . . . . . . . . . . . . . . . . . . . 60
Air Pollution and Asthma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60
Hospital Admission Studies for Asthmatic Patients . . . . . . . . . . . . 61
7.1
7.2
8.0
9.0
9.2
PLAUSIBLE MECHANISMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63
Biological Plausibility: Linkage between particulate matter (PM) and
adverse biological effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63
Inflammatory Response . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64
10.1
10.2
TOXICOLOGICAL EFFECTS OF DIESEL EMISSIONS . . . . . . . . . . 66
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66
Non-Cancer Toxicological Effects of Diesel Emissions . . . . . . . . 66
9.1
10.0
emission.
10.3
Cancer and Other Risks from Diesel Emissions . . . . . . . . . . . . . . . 73
11.1
11.2
CARCINOGENIC/GENOTOXIC POTENTIAL OF URBAN AIR . . 75
Epidemiology Studies for diesel emissions . . . . . . . . . . . . . . . . . . . 75
Extrapolation issues for lung cancer from rodent studies. . . . . . . . 77
11.0
12.0
REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78
emission.
Resource Document - Page 1
1.0
INTRODUCTION
Over the past decade, many epidemiologic studies using advanced statistical techniques have
shown an association between exposure to small, short-term increases in particulate matter levels
(PM) and increases in daily mortality and symptoms of certain illnesses (Dockery and Pope,
1994; Schwartz, 1994a; Katsouyanni et al., 1990, Samet et al., 2000b). For example, these studies
have shown an increase in mortality (death) due to respiratory and cardiovascular diseases and
a worsening of symptoms in people with asthma. Generally, in both short-term and long-term
studies, the magnitude of the effect of PM exposure was small- a great deal smaller than the
effects of tobacco smoke on disease that have been reported in similar epidemiologic studies.
Widespread exposure to particles may be associated with measurable effects on public health.
1.1
Causality
The assessment of the impact of air pollution inherently assumes that air pollution causes
increased morbidity (incidence or prevalence of disease(s) in a population) and mortality.
Although most air pollution experts may claim “causality” the level of evidence for such
associations with specific pollutants may vary for different health outcomes. In general, much
less research has been undertaken to assess the long-term effects of air pollution. Most impact
studies include some form of enumeration of hospital admissions, asthma attacks or acute
respiratory symptoms. The long-term effect of air pollution on death rates or years of life lost
and chronic bronchitis remain more controversial (Künzli, 2002). Studies such as one on
Seventh Day Adventists in California have attempted to evaluate the influence of ambient air
pollution on the severity of chronic bronchitis among smokers and non-smokers (Abbey et al.,
1999).
1.2
The Risk Function
The risk function for attributing adverse outcomes to air pollutants is generally dependent upon
the results of epidemiology studies.
The parameter of interest is often called
“exposure/response” or “dose/response” despite the fact that most epidemiological studies
directly measure neither exposure nor dose as these are formally described (Künzli, 2002).
The health impact of air pollution may be measured in a number of ways, but none as effectively
in the public’s perception as that of death. Two concepts of death must be distinguished. On
one side, death is an event, so monitoring death serves as a method of generating count data.
On the other hand, death is a delimiter of survival time or life expectancy.
Time-series studies provide an efficient means for investigating the association of the daily
number of death counts and the level of air pollution on that day or a few days before. On a global
scale these estimates are rather consistent (Samet et al., 2000). However, these time-series
studies have a restricted short-term horizon for effect which presents a difficulty for health
emission.
assessment of chronic exposure (Künzli, 2002). For example, a particular set of air pollutant
conditions might trigger or exacerbate a non-fatal respiratory condition or induce a myocardial
infarction. Neither of these may be immediately responsible for mortality, but could result in
mortality within a few weeks. Thus, time-series studies will not attribute such “late-onset death”
to an event because the vast majority of such studies only investigate the association of mortality
with nearly coincident conditions of ambient air quality.
In particular, time-series approaches do not assess life time lost, nor do they account for
shortening or reduction in life expectancy due to long-term cumulative exposure (Künzli, 2002).
It is possible that prolonged exposure to elevated levels of air pollution could lead to recurrent
infections that in the longer run could result in a chronic respiratory condition (Zemp et al.,
1999). Although the actual “event” of dying may not be related to the level of air pollution at
the time of death or to the period just preceding death, such premature death should clearly be
part of the air pollution health impact assessment. If such late onset of mortality is not
considered, the cases attributable to either a specific level of air pollution or to an annual average
will be underestimated.
Cohort studies directly assess the impact of cumulative air pollution experience on person-time
or time-to-death. While some very short-term, high exposure victims of air pollution may not
be captured in a cohort study approach, a more accurate assessment of overall effect of air
pollution should be provided by concentration/response functions derived from cohort studies
(Künzli et al., 2001). Risk functions based on the three well established cohort studies carried
out to date are five- to ten-times greater than those derived from time-series studies (Dockery
et al., 1993; Pope et al., 1995; Abbey et al., 1999).
Estimates for the annual years of life lost may be quantified by taking into account the age
structure of the exposed population. Assuming the cardiorespiratory causes of death, the age
of those for whom air pollution is attributed as a cause of death may be rather old. So far,
published studies do not provide estimates of the years of life lost nor the age structure of the
premature death, thus indirect estimates or assumptions must be relied upon, using life tables
of the population concerned (Künzli, 2002). The shortening in population mean life expectancy
has been estimated to be in the range of ~ 6 to >24 months for an increase in the annual mean
PM10 of 10 :g/m3 (Brunekreef, 1997; Pope, 2000; Künzli, 2002).
1.2.1
Reference level for air pollution: Impact above what level of
pollution?
The establishment of non-zero target values for clean air presumes that some health impact of
air pollution may be accepted (Künzli, 2002).
The choice of the air pollution reference value is of major influence for the number of cases
attributed to air pollution. The determination of such a level is important for the purpose of
setting public health policy. Policy makers need to set air quality regulations that are intended
to protect health, including the health of those most vulnerable to the effects of air pollutants.
emission.
To date research based on time series or cohort studies has failed to obtain evidence for a noeffect threshold. Thus, similar to carcinogens, the “natural threshold” might be zero exposure.
In the case of carcinogenic substances (whether natural or anthropogenic) environmental policy
has accepts the notion of diminimus risk (usually calculated incidence of less than one additional
cancer per 106).
In contrast to smoking where a reduction in smoking behaviour results in an immediate
reduction in exposure, individuals have very little influence on outdoor air quality. The health
risks of ambient air pollution must be considered very small risks on the individual level. Thus
individual contributions to improvements in air quality may have literally no personal health
benefit (Künzli, 2002). On the other hand, cooperative efforts to reduce emissions such as
through the general use of low a emission vehicles with high efficiency power sources by a large
segment of the population can lead to measurable health benefits.
The purpose of this document is to review recent research approaches and advances relating to
assessment of health effects of particulate matter in air. Step one of this process is the review
of research efforts reported in the publically available literature. These investigations are aimed
at addressing some of the key questions that link toxicological effects of individual substances,
and mixtures present, or likely to be present in air with identifiable impacts upon human health.
An excellent review of the physical characteristics and associated human health implications for
particulate matter is available through the U.S. EPA third draft review criteria document (2002a).
The type of research that has engaged the scientific effort of academic, government, as well as
industry associations, special interest groups, and other research institutions covers a wide range
to epidemiology and toxicology. The review of these efforts has been divided into the following
areas of interest:
Epidemiology of particulate matter and gaseous pollutants:
•
•
•
•
•
•
•
Time-series studies relating ambient pollution to health effects using
medical record data.
Mortality associations with urban air quality.
Morbidity/mortality due to respiratory disease with urban air quality.
Morbidity/mortality due to cardiac disease with urban air quality.
General associations of morbidity with mobile emissions and vehicular
traffic.
The interaction of urban air quality and respiratory asthma and allergy.
Morbidity of asthma and linkages to environmental factors present in
urban air.
Developing methodologies in personal exposure assessment for urban PM and air toxics:
•
•
•
Association of indicators for respiratory effects (lung function) with
urban air quality.
Lung function studies.
Chronic exposure to urban air in adults, children and groups with
emission.
special sensitivity.
Biological plausibility for respiratory effects:
•
•
Linkages to particulate matter.
Linkages with metals and reactive oxygen species.
Ultimately, if medical science hopes to be able to attribute adverse health effects to changing or
deteriorating urban air quality it will require tools that give answers as to the causes of the effects
observed. While much of the focus of the recent investigation on effects on populations has
been within urban boundaries, there is no reason to believe that health impacts are restricted to
cities. In fact, there is increasing evidence that for certain types of combustion emissions, no
clear evidence of a threshold has been observed. Analytical tools for predicting effects will be
developed as a result of an understanding of the relationship between exposure (to generally low
ambient levels of pollutants), and responses in sensitive receptors. In some areas, there have
been quite rapid advances in methodologies and the development of analytical instruments to
quantify effects and define biological responses observed as indicators or biomarkers of
exposure.
1.2.2
What Are the Sources of Airborne Particles?
The sources of PM are numerous...both naturally occurring processes and human activities
contribute to total ambient PM. Naturally occurring PM includes dust from the earth's surface
(crustal material), sea salt in coastal areas, and biologic material in the form of pollen, spores,
or plant and animal debris. In some rural areas, periodic forest fires produce large amounts of
PM. In urban environments, particles arise mainly as a result of combustion from mobile
sources such as cars, buses, ships, trucks, and construction equipment, and from stationary
sources such as heating furnaces, power plants, and factories. Near highways, motor vehicle
emissions may dominate the pollution mixture, but in other locations emissions from a power
plant or steel mill may be the main source of particulate pollutants.
A significant fraction of PM, referred to as secondary particles, is produced by chemical reactions
in the atmosphere; nitrogen oxides, sulfur dioxide, and organic compounds react with ozone and
other reactive molecules (including free radicals) to form nitrates, sulfates, and other particles.
People are also exposed to PM indoors, mostly from cigarette smoke, home heating sources
(such as woodburning stoves), and cooking, but also from outdoor PM sources that easily
penetrate the indoor environment. Indoor exposure may be substantial because this is where
most people spend the majority of their time. However, when outdoors, people tend to be
more active, and thus have higher respiration rates. People indulging in strenuous activity may
inhale a larger amount of pollutants because they inhale a larger amount of air in any given time
period than people who remain indoors. The relative attribution of inhaled pollutants and the
associated health impacts of indoor vs outdoor exposure to sources of PM is the subject of
continued scientific and medical investigation.
emission.
1.2.3
Health Effects of Particulate Matter
In humans, the primary organ directly affected by deteriorating or poor air quality is the lung.
The lung and its associated respiratory structures has defense mechanisms usually adequate to
deal with particles deposited on its surfaces. The volume of ambient air exposure in adults
ranges between 10,000 and 20,000 Litres per day. The lung surface area ranges from 40 to 120
m2 (Salvi and Holgate, 1999). Mechanisms available to the lung for management of particulate
matter include mechanical removal and biochemical neutralization. Larger particles that are
generally deposited in the conduction airways of the trachea and bronchi are removed by
mucociliary clearance. These particles are transported up the airways into the throat and finally
swallowed. In the oxygen absorbing area of the lung, particles are absorbed by scavenging
macrophages that then carry the engulfed or phagocytosed particles in one of two directions:
(1) up to the mucociliary clearance system; or (2) through the alveolar wall. Internal clearance
is achieved by transport through the lymphatic vessels into the lymph nodes.
Particles rarely enter the lung unaccompanied by gaseous pollutants such as ozone (O3), sulphur
dioxide (SO2), oxides of nitrogen (NO, NO2). Particles may include transition metals, which can
become reactive in biological systems to produce powerful oxidizing reactions. There are a
variety of mechanisms available to epithelial tissues of the lung to manage reactive oxygen
species, and to mediate damage to membrane structures (Salvi and Holgate, 1999). Among
patients compromised for lung function, small particle loadings may produce significant adverse
effects. Many reports have established associations of health effects with material in ambient
air, using indices such as mortality, acute asthma, chronic bronchitis, respiratory tract infections,
ischemic heart disease and strokes. Recent research into the toxicology of air particulate has
helped to understand the fundamental mechanisms of cytotoxicity induced from exposure of
tissues to ambient urban air pollutants.
Ambient particles have been associated with cardiovascular mortality and morbidity. It has been
suggested that ultrafine particles transported to alveolar tissue produce low grade inflammatory
responses. This might lead to increased coagulatability, and altered blood rheology (Salvi and
Holgate, 1999). Several hematological factors could be linked in this pathology, affecting
increased viscosity, fibrinogen factor VII and plasminogen activator inhibitor that occur because
of inflammatory action. These reactions are predictive of cardiovascular disease. For example,
a one hour exposure of human subjects to diesel particulate matter increased circulating platelets
and neutrophils in subjects. These changes were attended by acute cell mediated inflammatory
responses in the airways (Salvi et al., 1997).
A growing number of epidemiologic studies, frequently employing sophisticated statistical
analyses, have observed that acute effects of particulate air pollution are associated with
increased mortality (Samet et al., 1995; Samet et al., 1997; Levy et al., 2000). The prolonged and
severe pollution episodes that occurred in London, Belgium and the United States in the 1950s
that have long been cited to document the adverse health effect of urban air pollution. In
contrast to these acute episodes resulting from high levels of pollution, more recent analyses cite
current levels of ambient air pollutants as the possible cause of increased respiratory illness. In
emission.
these more recent reviews of the effects of air pollution on human health, the effects in people
suffering from cardiovascular and respiratory diseases have been noted at ambient air particulate
levels below current accepted air quality standards.
The renewed interest in the potential adverse effects of air pollution has been stimulated by the
hypothesis that long-term exposure to particulate air pollution may be responsible for an
increase in respiratory and/or cardiovascular disease in urban populations. Recent evidence
appears to support the view that the morbidity (adverse, non-lethal health effects) and mortality
associated with PM is attributable to a combination of pollutants that include ozone, sulfur
dioxide, oxides of nitrogen and carbon monoxide. Animal studies indicate that specific particle
types (nitrates, sulfates and sulfuric acid) impair pulmonary function, but have failed to support
the mortality findings of the epidemiology studies.
emission.
2.0
WHAT IS PARTICULATE MATTER?
2.1
What Are the Physical and Chemical Characteristics of
Particles?
PM is a complex mixture of solid and liquid particles. This mixture can vary greatly in size,
composition, and concentration. The components of the mixture of substances contained in
a specific sample of PM is dependent on the source(s) generating the particles as well as such
factors as geographic location, season, day, and even time of day. Ambient air particulate matter
contains particles of various sizes and composition. Anthropogenic sources of ambient particles
include mobile sources (engines powered by diesel, gasoline and jet fuels), stationary sources (oil
and gas fired boilers and electric utilities), and other sources (wood-burning fireplaces, paved
and unpaved roads, cigarette smoking and food preparation).
2.1.1
Size
For convenience, particulate matter is operationally defined by two categories currently used
various agencies for purposes of the regulation of ambient PM. These are PM2.5 and PM10,
which refer to particles with mean aerodynamic diameters smaller than 2.5 :m and 10 :m,
respectively. The size of ambient air particles ranges over a wide scale, from approximately
0.005 to 100 :m in aerodynamic diameter (that is, from the size of just a few atoms to about
the thickness of a human hair). Researchers have defined size categories of these particles
differently. Figure 1 shows that the distribution of particles measured in urban air falls into
three main modes based on their aerodynamic diameter: coarse mode (larger than 1 :m), and fine
mode composed of nuclei mode (smaller than about 0.1 :m), and accumulation mode (between
approximately 0.1 and 1 :m), and. The formal definitions of these particle distributions are
given below (U.S. EPA, 2002a).
Coarse Mode: The distribution of particles with diameters mostly greater than the minimum in
the particle mass or volume distributions, which generally occurs between 1 and 3 :m. These
particles are usually mechanically generated (e.g., from wind erosion of crustal material).
Fine Mode: The distribution of particles with diameters mostly smaller than the minimum in the
particle mass or volume distributions, which generally occurs between 1 and 3 :m. These
particles are generated in combustion or formed from gases. The fine mode includes the accumulation
mode and the nuclei mode.
Nuclei Mode: That portion of the fine particle mode with diameters below about 0.1 :m. The
nuclei mode can be observed as a separate mode in mass or volume distributions only in clean
or remote areas or near sources of new particle formation by nucleation. Toxicologists and
epidemiologists use ultrafine to refer to particles in the nuclei-mode size range. Aerosol
physicists and material scientists tend to use nanoparticles to refer to particles in this size range
generated in the laboratory.
emission.
Figure 1: Volume size distribution measured in traffic showing fine-mode and coarse-mode particles and
accumulation modes within the fine-particle mode. Geometric Mean Diameter by Volume (DGV) and
geometric standard deviation (Fg) are shown for each mode. Transformation mechanisms (nucleation,
condensation and coagulation) that lead to growth of particle size are also shown (Adapted from U.S. EPA,
2002a).
Accumulation Mode: That portion of the fine particle mode with diameters above about 0.1 :m.
Accumulation-mode particles normally do not grow into the coarse mode. Nuclei-mode
particles grow by coagulation (two particles combining to form one) or by condensation (lowequilibrium vapour pressure gas molecules condensing on a particle) and “accumulate” in this
size range.
Other definitions of these particles used in health effects studies and for regulation are: ultrafine
particles, smaller than about 0.1 :m in aerodynamic diameter (corresponding in size to nucleimode particles); and fine particles, smaller than 1 :m in aerodynamic diameter (containing all of
the nuclei-mode and accumulation-mode particles).
emission.
As illustrated in Figure 1, the largest particles (coarse particles in particular) form the highest
proportion of the mass of ambient particles; the smallest, ultrafine particles, comprise only 1%
to 8% of this mass. Coarse particles are generated mainly by mechanical processes that break
down material from a variety of non combustion-related sources into dust.
Ultrafine particles do not stay isolated for long periods the atmosphere, but tend to form fine
particles either by coagulating (two or more small particles combining) or condensing (gas
molecules condensing onto a solid particle). Fine and Ultrafine particles are always present in the
ambient air because of their generation from combustion sources. Ultrafine particles are present
in very high numbers. In a fixed volume ultrafine particles have greater total surface area than larger
particles.
2.1.2
Composition of Particulate Matter
A full physical characterization of particulate matter is available in the U.S. EPA Criteria
Document (U.S. EPA, 2002a). The composition of PM varies greatly and depends upon many
factors, including source, climate, and location. In the North America, for example, nitrates
tend to predominate in the west, whereas sulfates predominate in the east; in addition, sulfate
levels are higher in summer than in fall or winter. Even in a single location the composition of
PM can vary from year to year, season to season, day to day, and within a day.
The major components of PM are metals, organic compounds, material of biologic origin, ions
(that is, positively or negatively charged atoms), reactive gases, and the particle core (which is
frequently composed of pure, or elemental, carbon). Secondary particles (a major subcomponent of the ion fraction) are mostly composed of ammonium sulfate, ammonium nitrate,
and secondary organic compounds that are produced in the atmosphere via reactions of gases
with reactive organic compounds (see Table 1).
In general, the composition of larger particles differs from that of smaller particles. The coarse
particle fraction consists mainly of insoluble crust-derived minerals, biologic material (such as
pollen and bacteria), and sea salts. By contrast, the ultrafine and fine fractions are composed
mainly of particles with a carbon core that contains a variety of metals, secondary particles, and
hydrocarbons.
emission.
Table 1:
Chemical Components of Fine PM-Associated with
Biological Effectsa
Component
Subcomponents
Effects on Biological Systems
Metals
Fe, V, Ni, Cu, Pt, others
Trigger inflammatory responses, cause damage to
DNA; alter cell permeability through production of
reactive oxygen species (ROS) in tissues
Organic
Compounds
Adsorbed VOC and semivolatile organics
VOC can be mutagens/carcinogens, and can
induce allergic and asthmatic responses
Biogenic
Materials
Bacteria, endotoxins,
viruses, plant and animal
debris and fungal spores
Plant pollens can trigger allergic responses in
sensitized persons; viruses/bacteria induce immune
responses in airways
Ions
Sulphate (NH4), nitrates
and acidity (secondary
particles)
H2SO4 can impair upper airway clearance
mechanisms and irritate asthmatics. Acidity
changed bioavailibility of metals and adsorbed
compounds.
Particle Core
Carbonaceous material
(Diesel)
Organic carbon induces lung irritation, epithelial
cell proliferation, inflammatory responses, fibrosis
on prolonged exposure
a
Adapted from HEI, 2002
2.2
Ambient Particulate Matter
There are substantial differences in the chemical composition of fractions of PM of different
sizes collected from different locations across Canada or North America. The PM2.5 fraction
is composed of sulfates, nitrate, lead, cadmium, organic carbon and ammonium salts. The
coarse fraction of the PM10 is mainly comprised of natural and anthropogenic sources.
Windblown agricultural soil and dust from roads, or construction sites are examples of coarse
anthropogenic PM. Smaller particles typically have a more complex composition and are
generated from fossil fuel combustion in power plants, automobiles, industrial boilers,
residential heating and other combustion sources. Sulphate (SO42-) has repeatedly been shown
to be the single most abundant component of fine particles. In urban environments, organic
carbon compounds are responsible for much of the remaining fine particle mass (Environment
Canada, 1999).
Another distinction made for PM is between those emitted directly into the atmosphere
(primary particles) and those that are formed in the atmosphere by the coalescence of gaseous
emissions with particles (secondary particles). Primary particles are formed as the result of
physical processes and may be a characteristic component of highway PM from abrasion and
friction wearing away the road surface as well as tire wear. Secondary particles formed through
chemical reactions will also be present in roadway PM, since vehicles are sources of oxides of
emission.
nitrogen (NOx), sulfur dioxide (SO2), and volatile organic carbon (VOC). Primary particles that
comprise the coarse fraction settle out of the atmosphere and remain a locally, while secondary
particles may be transported some distance, thus effecting regional pollutant levels
(Environment Canada, 1999). This does not alter the pattern of direct emissions, but may be
a consideration for re-entrainment of particles in the disturbed air patterns created by moving
vehicles.
Ambient levels of PM are routinely monitored using a variety of technologies. None of these
are able to directly identify the source of PM with any degree of certainty, necessitating
laboratory analysis. It is important to recognize that the composition of PM collected or
monitored at a particular location may be quite varied because of the many different sources of
emissions that eventually contribute to the overall PM loading. The lack of a well defined
description for particulate matter in individual studies is a significant contributor to the
uncertainty in the evaluation of health risks. A recent study of sources of PM10 in Las Vegas
Nevada has determined that much of the particulate matter collected was from fugitive dust
emissions (Chow et al., 1999). The study, which employed a large array of monitoring sites,
found that the activities at distances up to, but not greater than 1 km could influence the
monitoring activities. The greater source contributor to ambient levels of PM was controlled
construction activities. Windblown dust sources arise from a wide variety of sources, and many
of these have low thresholds suspension velocities. Thus, “natural dust” from the surrounding
regions appeared to make little overall contribution to the levels of PM10 observed. Land uses
near sampling sites can significantly impact ambient measurements. Monitoring sites located
within 0.5 km of a source of emission could observe levels up to four times greater than might
be noted at nearby residential or commercial locations. Primary vehicle exhaust was the second
largest contributor of PM10 emissions associated with traffic movement (Chow et al., 1999).
Paved road dust as well as construction related PM fluctuate in their relative contribution to
ambient measurements. In some cases, the use of a single monitor to represent a large
population may be subject to localized sources, and therefore be unduly influenced by location
within an urban area or changing patterns of land use.
emission.
3.0
PM AND RESPIRATORY HEALTH RESPONSES
3.1
The Respiratory Tract as a Target
The respiratory tract is the prime target for effects of inhaled PM. The respiratory tract is
physiologically complex. The functional role of the respiratory tract is to move gases from the
nose or mouth via the airways to the alveoli where the exchange of the gases oxygen into and
carbon dioxide out of the body occurs. When the inhaled breath is accompanied by PM, a
number of associated biological functions are also effected. These functions include its role as
an elaborate biochemical system that aids in olfaction and in detoxification (decreasing toxic
activity) or activation (increasing toxic activity) of inhaled chemicals. The respiratory tract also
has immunological capabilities, serves as a blood reservoir, and plays a key role in acid-base
balance (McClellan and Miller, 1997). The principal mechanisms of deposition of inhaled PM
are impaction, sedimentation, and diffusion. These three processes are strongly influenced by
the aerodynamic characteristics of PM.
•
•
•
Impaction is the process by which particles suspended in the air travel along
their original path due to their inertia and impact on a surface rather than
following the airflow around a bend, as might occur near an airway
bifurcation.
Sedimentation (settling out) of particles predominates when air flow is low.
The rate of sedimentation is determined by the terminal velocity of particles.
As particles increase in aerodynamic diameter, the terminal velocity
increases. This increases the role of sedimentation in deposition of larger
particles.
Diffusion is the major process governing deposition of particles less than
0.5 :m in diameter. Diffusional displacement motion increases for particles
of smaller size. Diffusional deposition is high in the pulmonary region,
where air flow is low.
There is a relatively large data base on experimental deposition studies in human subjects (U.S.
EPA, 1996). While the generalized values for rates of deposition can be described based on
laboratory data, inter-subject variability is quite significant, particularly for tracheobronchial and
pulmonary deposition of particles 2 to 6 :m in aerodynamic diameter. Particle deposition can
vary four- to five-fold among subjects for a given particle size and spontaneous breathing
pattern. This variation makes it extremely difficult to establish the specific responses and the
magnitude of these responses one can expect in individuals exposed to a given level of PM.
3.1.1
Exposure to Particles and Effects on Cells and Tissues
Multiple biological responses sequentially act to prevent or reduce the effects of exposure to
particulate pollution. The first important level is a barrier of cells and fluids that acts to
physically exclude the intrusion of toxic substances. Fluid secretions, such as mucus lining the
airways, and ciliated cells are important elements of the system that is responsible for the
emission.
exclusion of particles, especially larger ones, from airways. Coughing, activated by particles
interacting with receptors on nerve cells in the airways, also helps to remove particles.
The second line of defence relies on scavenging cells that ingest foreign material into vacuoles,
release digestive enzymes, and attempt to destroy it. Important scavenger cells include
macrophages and neutrophils. On occasion, the proteins released by macrophages and
neutrophils trigger additional responses such as stimulating or attracting lymphocytes, effectively
widening the level of response.
Active engulfment of particles deposited on epithelial cell surfaces in the airways by resident
macrophages and neutrophils is termed cell activation. Activated cells are known to synthesize
and release potent compounds referred to as reactive oxygen species (ROS) including hydrogen
peroxide. Subsequent reactions include the production of cytokines (proteins) and smaller
molecules (chemokines) that are secreted into affected areas. Cytokines and chemokines are
molecules that act as mediators of a cascade of cellular responses. They interact with specific
receptors on the surfaces of many cell types, causing these target cells to be activated in turn.
This recruitment process induces cells to leave the bloodstream and enter the interstitial fluid
spaces of the airways. Exposure to PM, therefore, can result in a characteristic inflammatory
response in surrounding tissue structures.
Inflammatory responses can be destructive, leading to epithelial cell damage and physical
deterioration of tissue integrity and cellular defences. This is sometimes reflected in an impaired
immune response to immunological challenges such as bacterial or microbial infection. At least
one of the cytokines produced by the inflammatory response in the airways stimulates the liver
to secrete a set of molecules known as acute-phase reactants. These molecules, which include
C-reactive protein and fibrinogen, appear in the circulation within 6 to 24 hours. Fibrinogen
binds to platelets and contributes to their aggregation. This can result in multiple effects
throughout the cardiovascular system, including an enhanced ability of the blood to clot
(increased coagulability). This may be important in the development of clots and myocardial
infarction, and thus the apparent association of increased fine PM and an elevated frequency of
heart attack.
Although cytokines may have a primary role in inducing these nonpulmonary effects, some
recent research also suggests that particles (and ultrafine particles in particular) or particle
components may physically move out of the airways and rapidly into the bloodstream to trigger
effects at distant sites. Deposition of particles in the airways can stimulate nerve cells in the
underlying tissue as well. Activation of these cells has been suggested to lead to changes in the
nervous system's control of the pattern of breathing, the heart rate, and heart rate variability (a
measure of the fluctuations in heart rate that occur in all individuals), and to affect other cardiac
electrophysiologic parameters. Thus, particle deposition in the airways can set off a cascade of
events in many different cells, potentially resulting in changes in tissues and organs at sites
progressively further away from the initial stimulus. These defence mechanisms are normal
responses in healthy individuals, but they may lead to deleterious changes in the host. Such
changes may be rapid and temporary and may resolve quickly; but depending on the level and
pattern of exposure and the agent to which the host is exposed, the changes may last longer.
emission.
It is not clear whether or how such changes are relevant to the development of PM-induced
adverse health effects at low levels of exposure. These changes are thought to have a greater
impact on individuals whose airway, cardiac, or vascular tissues have been previously damaged
(HEI, 2002).
A scheme representing key steps observed during the activation of inflammatory response to
inhaled particulate matter is presented in Figure 2
emission.
PARTICLES DEPOSITED
IN AIRWAYS
Neural pathways
Autonomic NS
Circulation & tissues
outside airways
Activation of cells
in the airway
Leukocyte activation
and inflammation
Cardiac
electrophysiologic
changes (ECG)
INFLAMMATION
Induction of acute-phase response
(C-reactive Protein)
Changes in
respiratory function
& increased fibrinogen
& blood coagulability
Thrombosis
CARDIAC EVENTS:
MYOCARDIAL INFARCTION,
ARRYTHMIA AND/OR DEATH
Figure 2:
Mechanisms and effects of fine particulate matter on airways and the cardiovascular system.
ECG is electrocardiogram. (Adapted from HEI, 2002)
emission.
3.1.2
Attributes of Particles That Are Associated with Toxicity
Studies have investigated the physical aspects (particularly size) and the chemical composition
of particles that induce effects in humans and other species. Several recent epidemiologic
studies (Schwartz et al., 1996; Fairley, 1999; Burnett et al., 2000; Ostro et al., 2000; Castillejos et
al., 2000; Hoek et al., 2000; Gwynn et al., 2000) in different locations have reported associations
between various health effects and different sizes and/or chemical components of the particles
to which the study populations had been exposed. Findings differed from study to study,
generally depending on where and when the study was conducted. For example, some
epidemiologic studies in Mexico City and the western United States have found health effects
associated with the coarse particle fraction (Ostro et al., 2000; Castillejos et al., 2000), but studies
conducted in other parts of the United States and in Canada have reported that effects of fine
particles predominate (Schwartz et al., 1996; Fairley, 1999; Burnett et al., 2000). In a recent study
in Germany, levels of both fine and ultrafine particles were associated with increased mortality
(Wichmann et al., 2000).
Several reasons may be suggested for such discrepancies among studies (HEI, 2000):
•
•
•
•
The nature of PM varies in different regions with different sources, climate,
and topography;
Studies use a variety of statistical methods to assess results;
Studies may use different measurements of PM (eg, PM2.5 vs ultrafine
particles); and
Studies may have different health endpoints.
The National Morbidity, Mortality, and Air Pollution Study (NMMAPS) (funded by the Health
Effects Institute) directed by Jonathan Samet, investigated the association of PM10 with mortality
using a unified method in the 90 largest US cities (Samet et al., 2000b). This study was designed
to address many of the criticisms raised about the time-series studies that had shown an
association between short-term increases in PM levels and increases in the number of deaths.
NMMAPS found that a 10 :g/m3 increase in PM10 resulted in an average increase of about
0.5% in mortality from all causes. Adding other pollutants to the model did not appear to affect
the result found with PM10. Regional differences in the PM10 effect were seen, with the largest
effects evident in the northeastern United States. One explanation for the regional differences
may be variation in the nature of PM; because it is a complex mixture, the same mass of PM10
in different places may include very different amounts of fine or ultrafine particles and particles
with different composition (eg, higher presence of some metals).
3.2
Establishing a Dose-Response Relationship for Effects of
Particulate Matter.
The respiratory tract is a complex organ. Alterations to physiological challenges include
physiological adaptation to mild dysfunction. In addition to responses involving the respiratory
system alone, the close integration of pulmonary and cardiac function raises the possibility
emission.
serious, life-threatening disease due to combined cardio-pulmonary effects.
Examination of dose metrics based on particle number for their toxicological significance has
been suggested (Maynard and Maynard, 2002) since significant species differences are predicted
to occur for fine-mode particles in the range of 0.1 to 0.3 :m. Individuals with compromised
lung status such as patients with chronic obstructive pulmonary disease (COPD) have fewer
functional alveolus than normal individuals. A reasonable argument might be made that
inhaling high concentrations of particles with diameters 0.1 to 0.3 :m may result in a localized
overloading of pulmonary clearance mechanisms for such individuals. This overloading, in turn,
could trigger a cascade of inflammatory-related events leading to acute morbidity, mortality, or
both (Godleski et al., 2000).
PM deposited in the respiratory tract may elicit a wide range of responses. PM may interact with
neurogenic receptors at different locations in the respiratory tract to alter breathing frequency
and depth. Sensory nerves can trigger coughing in response to a range of inhaled agents
including chemical irritants and dusts.
Wichmann et al. (2000) characterized the sizes of particles in the ambient air of Erfurt, Germany,
and determined whether they were related to changes in daily mortality. They reported that over
a three-year period the concentrations of both ultrafine (PM< 0.1) and fine particles (PM0.1-2.5)
were associated with increased daily mortality. These findings provided the first evidence that
ultrafine particles were associated with human mortality, but did not indicate whether ultrafine
particles were more toxic than larger particles. Lippmann et al. (2000) compared day-to-day
fluctuations in hospital admissions of older people and deaths in the Detroit-Windsor area with
day-to-day fluctuations in levels of different ambient PM size fractions. They found that four
of the five size fractions they evaluated were associated with increased morbidity and mortality.
These were total suspended particles (TSP), PM10, PM2.5–10, and PM2.5. The magnitude of the
association was similar for all four fractions. Since all fractions contained PM2.5 it could be
concluded that the fine particulate was the fraction of greatest health impact. The largest
particle size fraction (between 10 :m and about 40 :m) was not associated with increased
morbidity and mortality. The investigators also reported that the particles fractionated by size
were more significantly associated with health outcomes than were the two chemical
components of ambient PM, acidity and sulfate, evaluated in the study.
Ambient air also contains many different organic compounds associated with combustion
particles. Some studies have shown that an organic fraction extracted from diesel exhaust
particles, an emission reported to enhance the induction of at least some of the characteristics
of the allergic response in humans and other species, enhances the synthesis of immunoglobulin
E (a key mediator of the allergic response) in vitro (Tenaka et al., 1995; Tsien et al., 1997). An
organic extract of diesel exhaust particles has been reported to have cytotoxic effects in
macrophages and epithelial cells in vitro (Nel et al., 2001).
Some experimental and epidemiologic studies have tried to associate health effects not only with
specific components of PM but also with specific sources of particles (Ostro et al., 2000;
Godleski et al., 2000; Laden et al., 2001). The statistical approaches in these studies, which
emission.
included factor analysis and principal component analysis, are based on assumptions about the
groups of elements that characterize an emission source.
3.3
How Do Particles Cause Effects?
Recent PM research has begun to identify plausible biological mechanisms that explain the
epidemiologic findings of associations between increased exposure to PM and increased
mortality. Laboratory studies initiated by Godleski et al. (2000) have used particles comparable
to or derived from those found in ambient air (concentrated ambient particles). Other
investigations used populations particularly susceptible to the effects of PM. These findings
present a credible view of how even low-level exposure to PM may alter the cardiovascular and
pulmonary systems and pose a particular threat to people with cardiovascular or respiratory
conditions. As shown diagrammatically in Figure 2, there appears to be increasing support for
a pattern of interrelated effects and mechanisms linking PM to cardiovascular and other specific
health outcomes. These include the following:
•
•
•
the induction of inflammatory responses in the airways;
the induction of systemic inflammatory and other vascular responses; and
changes in neural control of heart function.
The major mechanisms by which PM induces toxicity (described in Figure 2, above) that have
been identified are discussed below.
3.3.1
Inflammatory Responses in the Airways
Recent controlled-exposure studies in humans indicate that different types of particles can
induce an inflammatory response in the airways, the site at which particles first deposit (Ghio
et al., 2000). This is measurable in a number of ways, including an increase in neutrophil number
and in levels of cytokines and chemokines associated with the inflammatory response. Salvi et
al. (1999, 2000) found airway inflammation in healthy subjects after exposure to diesel exhaust.
Nordenhäll et al. (2001) have shown that when asthmatics are exposed to high concentrations
of diesel exhaust, typical hyperresponsiveness of the respiratory system is detectable in small
airways. Although impacts on lung function were not immediately observed , a delayed
sensitivity to asthmatic/respiratory challenge was detected. During the post diesel exposure
period, inflammatory responses including cytokine IL-6 production was initiated. No other
statistically significant induction of a variety of mediators of the inflammatory response were
detected in subjects post-diesel exposure (Nordenhäll et al., 2001).
emission.
In animals and from in vitro experimental results it has been demonstrated that the metal and
organic components of PM can induce inflammatory cytokine and chemokine formation as the
end result of oxidative stress pathways inside cells. These pathways generate what are known as
reactive oxygen species, including free radicals, hydrogen peroxide, and superoxide (Nel et al.,
2001; Donaldson et al., 1997). The induction of an inflammatory response by PM in the airways
may damage not only the epithelial cell layer at the surface of the tissue but also other airway
cells such as macrophages. Renwick et al. (2001) have shown that fine and ultrafine particles
decreased phagocytosis in a macrophage cell.
3.3.2
Systemic Inflammatory and Other Vascular Responses
Systemic responses to PM exposure include inflammatory effects within hours after exposure.
In some studies, either particles per se (and ultrafine particles in particular) or components that
may detach or dissolve from particles may move rapidly into the bloodstream and reach other
tissues (Nemmar et al., 2001, 2002). One marker of systemic inflammation that has been
detected after exposure to PM is an increased number of circulating neutrophils in humans
(Salvi et al., 1999), in rats (Gordon et al., 1998), and in dogs (Clarke et al., 2000). Increased bone
marrow production of immature neutrophils has also been reported (Terashima et al., 1997;
Suwa et al., 2002).
Epidemiologic studies have described associations between PM exposure and other vascular
factors, and controlled-exposure studies have reported PM-dependent effects on levels of
additional vascular factors (Peters et al., 1997c; Seaton et al., 1999; Pekkanen et al., 2000;
Schwartz, 2001; Vincent et al., 2001; Peters et al., 2001a; Ibald-Mulli et al., 2001). These factors
include fibrinogen, plasma viscosity, platelet numbers, C-reactive protein, endothelin levels, and
blood pressure. Several of these factors (fibrinogen, C-reactive protein, and blood pressure) are
independently associated with increased risk of cardiovascular disease, which could affect
susceptibility to the acute effects of PM (HEI, 2002).
The changes in vascular parameters that occur after particle exposure suggest that exposure to
PM may lead to higher levels of fibrinogen, in turn increasing plasma viscosity, and the ability
of blood to coagulate (Godleski et al., 2000). This may result in an increased tendency to form
clots and thrombi (aggregations of platelets and other blood components causing vascular
obstruction). It can be assumed that inducing clots or thrombi in persons with previously
damaged cardiac or vascular systems could lead to additional risk including death. Individuals
with atherosclerosis may be particularly at risk. Atherosclerosis is characterized by a thickening
and hardening of the arteries in which plaque (deposits of cholesterol and other fats, plus fibrin
and inflammatory cells and factors) narrows the arteries and decreases the arterial blood flow.
In atherosclerosis, the functions of endothelial cells, the cells lining the blood vessel, are also
impaired. This results in additional production of mediators that promote vasoconstriction, the
narrowing of blood vessels (HEI, 2002; Moyer et al., 2002.).
Blockage of a coronary artery can result in reduction or loss of the supply of oxygen to the heart
muscle. This condition, myocardial ischemia, may lead to heart damage and arrhythmias,
emission.
disturbances in the rhythmic beating of the heart. Arrhythmias, such as ventricular fibrillation,
may have serious and potentially fatal consequences because they can lead to a heart attack
(myocardial infarction). Arrhythmias may also develop as a consequence of changes in the
neural control of heart function. Several studies in humans and other species link exposure to
PM with changes in cardiac function, including inducing arrhythmias and increasing the
incidence of myocardial infarctions (Watkinson et al., 1998; Poloniecki et al., 1997; Peters et al.,
2000a; Campen et al., 2000; Campen et al., 2001). Short-term exposure to PM2.5 has been linked
to the fairly rapid incidence of myocardial infarction (Peters et al., 2001b). Godleski et al. (2000)
induced a temporary coronary occlusion (cutting off blood supply to the heart via the coronary
artery) in a small number of dogs and then exposed them to concentrated air particulate sourced
from the ambient environment. They found that these exposures induced a more rapid and
larger elevation in the ST segment on an electrocardiogram (ECG) than did exposure to particlefree air. This change in the ST segment is one of the characteristic signs of the onset of
myocardial ischemia (Allred et al., 1991). Although only one study, this finding supports the
mechanism by which people with atherosclerotic arteries may be more vulnerable to cardiac
problems, such as fatal arrhythmias, when exposed to PM. Kodavanti et al. (2000) have reported
that spontaneously hypertensive rats exposed to residual fly ash also show enhanced changes
in the ST segment of an ECG.
In a strain of rabbit prone to the development of atherosclerosis, short-term PM10 exposure
induced atherosclerotic lesions to progress to more advanced stages, potentially making the host
more vulnerable to an acute coronary event (Gordon et al., 1998). Confirmation of the relevance
of this observation to other species is required, but these findings could supply the mechanistic
link between PM exposure and increased incidence of myocardial infarction (HEI, 2002). Other
vascular parameters affected by PM exposure in experimental animals have been reported by
Vincent et al. (2001). Exposure to very high levels of three types of particles increased blood
levels of endothelins: molecules that affect blood pressure by inducing vasoconstriction (Vincent
et al., 2001). Particles used in the procedure included (1) concentrates from ambient Ottawa air
gathered and resuspended in air for the laboratory exposure chamber; (2) a washed particulate
sample of resuspended Ottawa air particles from which soluble constituents had been removed;
and (3) resuspended diesel soot. The resuspended Ottawa particles also caused a small increase
in blood pressure, but diesel soot (composed predominantly of carbon) did not. The removal
of metals through washing also eliminated the effect on blood pressure, indicating the greater
importance of metals in the production of the response (Adamson et al., 2000).
3.3
Neural Control of Heart Function
Data from recent studies suggest that PM exposure can affect the neural control of heart
function. Studies of older people (>65 yrs), or in people suffering from heart disease, indicate
that PM exposure was associated with decreased heart rate variability (Liao et al., 1999; Pope et
al., 1999b; Gold et al., 2000). Variability and nervous control of heart rate depends on the
autonomic nerves, and a balance between the sympathetic and parasympathetic components of
the nervous system (HEI, 2002). Although reduced heart rate variability is associated with
worse outcome in individuals with existing cardiac disease, the clinical significance of similar
emission.
decreases in healthy individuals is unknown (HEI, 2002).
3.4
Mechanisms of Particle Effects: Conclusions
Results from epidemiologic and experimental studies since the mid 1990s have:
•
•
•
Started to identify characteristics of particles that may induce health effects.
Suggested plausible biologic mechanisms that may underlie the reported associations
between short-term increases in PM levels and increases in morbidity and mortality.
Identified certain groups in the population who appear to be at increased risk from
exposure to PM.
A credible pathophysiologic basis has been emerging to explain how increases in PM levels
might increase morbidity and mortality. Multiple pathways have been identified, but additional
research is required establish the relevance of both the pathways, and of the endpoints
measured, to the induction of health effects at exposure levels similar to ambient levels (HEI,
2002).
Modelling the effects of chemical exposure to such complex mixtures has always been
problematic for toxicologists. While progress has been made in identifying the potentially toxic
effects of individual components of the PM mixture, the critical questions linking size and
chemical composition of the components to toxic effects remain only partially answered.
A more complete picture of the cardiac, pulmonary, and vascular effects of PM exposure is
emerging. There remains considerable uncertainty regarding the exact role of PM, however,
because the reported results are not always consistent from study to study. For example,
fibrinogen levels were increased in a human controlled exposure to concentrated particles
(CAPs) (Ghio et al., 2000). Increased levels of fibrinogen were positively associated with PM
exposures evaluated in an epidemiologic study reported by Pekkanen et al., (2000), but had a
negative association with PM10 exposure in another epidemiologic study (Seaton et al., 1999).
In addition, the relation between PM and health varies depending upon the investigators' choice
of different “lags” in time between PM exposure on a particular day and the day of the observed
health endpoint (Wichmann et al., 2000; Lippmann et al., 2000). Checkoway et al. (2000) found
no link between sudden cardiac death and PM exposure on the day of death or for exposures
likely to have occurred up to five days before death. Since this study was conducted in people
with no known heart disease, the conclusion was that for people with no evidence of heart
disease there is little or no increased risk of sudden cardiac death as a result of exposure to
ambient levels of particulate matter.
emission.
3.4.1
Are Some Individuals or Groups Particularly Susceptible to
the Effects of Particles?
Many epidemiologic and toxicologic studies have compared the effects of PM on morbidity or
mortality among different groups. Several studies have shown that estimates of short-term PM
effects on acute mortality are increased for people with cardiovascular disease (Schwartz, 1993;
Schwartz et al, 1996; Ostro et al., 2000; Samet et al., 2000a; Samet et al., 2000b), preexisting
respiratory conditions (Schwartz, 1994b; Goldberg et al., 2000), and older people with
preexisting respiratory or cardiovascular disease (Kelsall et al., 1997; Ostro et al., 1996; Simpson
et al., 1997). Goldberg et al. (2000) used the extensive health records compiled by the Quebec
government on hospital admissions and doctor visits to evaluate the underlying causes of
mortality associated with PM exposure. In the Montreal area, three daily measures of PM level
(coefficient of haze, and levels of sulfate and PM2.5) were associated with mortality from acute
lower, but not upper, respiratory disease, any cardiovascular disease, and other non-accidental
causes of death, including diabetes. This supports the hypothesis that people with respiratory
or cardiovascular disease are at increased risk from PM exposure.
Other epidemiologic studies have suggested that ambient PM may affect pregnant women and
their fetuses and infants. The results include increases in low birth weight and more infants
born prematurely (Woodruff et al., 1997; Dejmek et al., 1999), and an increase in infant and child
mortality (Loomis et al., 1999; Ostro et al., 1999b; Ritz et al., 2000). Some controlled-exposure
studies in animals, particularly those with characteristics that mimic certain human cardiac and
pulmonary conditions, support the idea that some groups are more sensitive to the effects of
PM than others.
emission.
4.0
LINKAGE OF HEALTH IMPACTS WITH
ENVIRONMENTAL POLLUTANTS THROUGH
EPIDEMIOLOGIC STUDIES.
4.1
Introduction
Secondary particle formation and the role of precursor gases reveals that there are links between
PM and a number of critical pollutants. Emissions of SO2 and NOx link the PM issue to acid
deposition and ground level ozone. PM is also linked to ozone through the common precursors
of NOx and VOC (Environment Canada, 1999). Confirmation of links between specific
pollutants, or mixtures of pollution emissions and human health endpoints has focussed science
on several types of approach, including epidemiological studies.
Epidemiological studies seek to determine whether statistical associations can be made between
human responses measured as, for example, admission to hospital for specific respiratory or
cardiovascular symptoms, and the level of pollution over some defined period. These are
subject to confounding factors such as prior exposure to other sources of pollution (smoking)
occupational history, temperature and the general health of the population. The complexity of
the associations among so many factors and the uncertainties associated with the available
scientific research make specific characterization of the health effects of PM10 and PM2.5 difficult
to quantify at this time. The primary considerations that need to be kept in mind are as follows:
1.
Exposure is widespread, but the risks of developing adverse health effects are very
low, necessitating the study of very large population groups in order to achieve
adequate statistical power.
2.
The adverse health effects that could be associated with air pollution have many
other etiologies.
3.
A complex relationship exists between weather, pollution and health effects and no
one pollutant can be concluded to contribute more than others to health effects.
Pollutant concentrations are often strongly correlated with each other and with
weather and seasonal parameters, making it impossible to determine with accuracy
the individual contributions of specific parameters to adverse health effects.
4.
Various different ways of measuring particulate matter (e.g., total suspended
particulate matter (TSP), PM10, PM2.5, black smoke, etc.) were used in the various
studies, as governed by the availability of monitoring. Nevertheless, adverse health
effects attributed to ambient particulate matter have consistently been demonstrated
by many researchers on different continents in separate epidemiological studies.
There is a coherence of the data with the health effects over a wide range of
exposures (Bates, 2000)
5.
Some of the underlying biological mechanisms of toxicity of ambient particulate
emission.
matter are becoming better understood.
6.
Exposure may be missclassified, since the pollution monitors may not reflect
exposure of individuals at sites distant from the monitors, and indoors, where
people spend the majority of their time. Therefore exposure assessments that rely
on aggregate data from fixed pollution monitoring stations may not reflect the actual
exposures of the individual subjects in the study.
While uncertainties associated with the current state of knowledge regarding toxic effects of
ambient PM2.5 and PM10 remain, the large volume of studies included in statistical analyses of
ninety cities in the United States based on hospital records show a clear relationship between
particulate matter, morbidity and mortality (Samet et al., 2000a, b).
4.2
Hospital Admission Studies
Hospitalizations for respiratory illness are one readily quantifiable measure of the respiratory
morbidity of a community during a specified time frame. Overall, the database on hospital
admissions provides support for the conclusions of the mortality studies. That is, that air
pollution is causally related to adverse cardiorespiratory health effects, but beyond a strong
association with PM, these effects cannot be attributed to any one component of the pollutant
mix.
There are positive associations between short-term concentrations of PM and hospital
admissions for respiratory-related and cardiac diseases. The magnitude of observed effects are
small but significant. The increases in mortality are limited to the elderly. For PM exposure in
this sub-population, any particular series of events that result in an increase in observed 24-hr
average concentration of PM10 of 10 :g/m3 (this includes PM2.5) results in an increased mortality
on the order of 0.5% to 1.7% (Environment Canada, 1999). Slightly larger effects have been
associated with effects such as exacerbation of asthma, loss of lung function and increased
respiratory irritability. These latter effects occur in all age groups, but asthma is a particularly
sensitive measure for effects in the young.
A recent report from the Health Effects Institute (HEI) called the National Morbidity,
Mortality, and Air Pollution Study (NMMAPS) (Samet et al., 2000b) has examined multiple
locations within the United States in order to characterize the effects of airborne particles less
than 10 :m in aerodynamic diameter (PM10) alone and in combination with gaseous air
pollutants in a consistent way. The investigators first estimated risk in each of twenty cities
using the same method, and then performed additional statistical analysis that combined these
results systematically. The result was that the investigators were able to recast the data, and to
examine health effects that were relevant to a greater population with a power of association
greater than any single city could provide. Within the United States, the 20 and 90 largest cities
were analysed for effects of PM10 and other pollutants on mortality. Samet et al. (2000) applied
a unified statistical method to 14 cities with daily PM10 data to examine effects on hospitalization
among those 65 years of age or older.
emission.
4.2.1
Assessment of Deaths Attributable to Air Pollution: LongTerm Predictions from Time-Series Studies
Impact assessment studies follow at least three different strategies: the exposure-response
function (slope) for mortality is based on either 1) time-series studies, 2) cohort studies, or 3)
an average estimate of time-series and cohort study results (Künzli et al., 2001). Air pollution
studies suggest that the greater the long-term cumulative exposure to ambient air pollution, the
more incidence (Abbey et al., 1995) or prevalence (Zemp et al., 1999) of chronic bronchitis
increases and lung function deteriorates (Ackermann-Liebrich et al., 1997). Coherent with these
effects on morbidity, long-term air pollution levels are associated with shorter times to death
(Dockery et al., 1993; Pope et al., 1995; Abbey et al., 1999). Time-series studies do not provide
sufficient information to evaluate long-term effects of air pollution (McMichael et al., 1998;
Künzli et al., 2001). Conceptually, time-series-based numbers of attributable cases of mortality associated with
pollution events must underestimate the total effect of air pollution.
4.2.2
The concept of death in time-series and cohort studies
In the field of air pollution epidemiology, nonviolent death is the key measure of interest. The
probability of death increases with increasing frailty or susceptibility. The “frailty” concept
(Zeger et al., 1999) is useful in acknowledging that, in most cases, the probability of death is
influenced not by one single factor (e.g. air pollution) but rather by a function of a whole set of
underlying conditions or risk factors. For example, preexisting diseases, genetic factors, age,
socioeconomic status, nutrition, and other environmental stressors may contribute to a person’s
frailty level (Künzli et al., 2001).
As one of the primary outcomes of interest used as a measure of effect in epidemiology studies,
the occurrence of death from air pollution may be influenced by factors that act shortly before
death. For example, an acute episode of pneumonia (linked in time with a pollution event) on
top of underlying chronic bronchitis (an increased frailty level) may be considered the terminal
cause or “exposure” that leads to mortality. There is evidence that the level of air pollution in
the days shortly before death is associated with the probability of dying (Samet et al., 2000a;
2000b). Schwartz has used analysis of time series data to show that exposure in the few days
prior to death as well as average concentrations in the weeks preceding death are associated with
the increased probability of death (Schwartz, 2000).
“Time to death” or survival time is also used as a measure of the impact of pollution (Künzli
et al., 2001). The concept of time lost (premature mortality) because of an exposure is a direct
measure of health impact of pollution, and a primary concern of the public. Time-series
analyses result in the suggestion that some members of an exposed population will die as a result
of exposure during a specific time period. This type of analysis does not account for prior or
preexisting health state of individuals, nor does it reflect the history of exposure(s) prior to the
event or time period of interest. There are a variety of well known “exposures,” including
morbidities, which have an impact on survival time (smoking, dietary habits, suffering due to
emission.
chronic bronchitis, reduced forced vital capacity, occupational exposures, etc.) (Künzli et al.,
2001). Several patterns of exposure may influence time to death, such as:
• a single past exposure that occurred over a short period (e.g., an accidental spill);
• repeated exposures of short duration;
• continuous exposure over a longer period of time;
• a pattern of short-term exposure just before death; or
• a combination of these events.
4.2.3
Four categories of air pollution-attributable deaths
Four different categories of air pollution-attributable deaths can be defined based on probable
patterns of exposure (i.e., acute exposure shortly before death, or short-term exposure versus
cumulative or long-term exposure). These categories have been defined by Künzli et al. (2001):
I. Air pollution increases both the risk of underlying diseases leading to frailty and the
short-term risk of death among the frail;
II. Air pollution increases the risk of chronic diseases leading to frailty but is unrelated
to timing of death;
III. Air pollution is unrelated to risk of chronic diseases but short-term exposure
increases mortality among persons who are frail; and
IV. Neither underlying chronic disease nor the event of death is related to exposure to
air pollution.
For example deaths in category I might occur among patients whose chronic bronchitis is
enhanced by historical long-term air pollution exposure. Such susceptible individuals may be
hospitalized with an acute, air pollution-related exacerbation of their illness (Wichmann et al.,
2000), leading to death shortly after admission.
For deaths assigned to category II, long-term or chronic exposure to elevated levels of pollution
might contribute to frailty or a reduced state of health, but the event or the occurrence of death
itself is unrelated to the levels of air pollution shortly before death. A person suffering from
chronic bronchitis may could die of acute pneumonia acquired during a clean air period. In this
instance it is probable that long-term cumulative exposure to air pollution contributed to
shortening of survival time, whereas air pollution during the final days of life had no further lifeshortening effect.
Death attributed to category III is not related to a reduced state of health because of air
pollution, but ambient air pollution experienced before death may trigger the terminal event.
For example, a person with diabetes mellitus may be susceptible to heart attacks due to longstanding coronary disease; in such a case, an air pollution episode may trigger the fatal infarction
leading to hospital admission (Schwartz, 1999), arrhythmia (Peters et al., 2000), or death
(Schwartz, 2000).
emission.
Finally, in category IV (sometimes referred to as ‘all cause non-traumatic mortality’), neither
disease history nor the event of death may be related to air pollution. Thus, deaths in category
IV will be attributed not to air pollution but to other causes only.
4.2.4
Air pollution attributable death in time-series studies
Time-series studies model the association between the probability of death and the level of air
pollution shortly before the death (the event). Time-series studies use counts as the outcome
measure (what number of events were recognized on a specific day under a given ambient
pollutant level?). This approach makes use of the temporal variability of air pollution to estimate
the mortality in an urban area due to air pollution. The duration of “exposure” cannot be years
or a lifetime, as temporal variability is a key requirement in the time-series approach; the longterm exposure history does not change from day to day. The time-series approach studies the
short-term relation between exposure and an event; that is, the time interval between the
measure of exposure used in the statistical model and the event (death) is short. A vast majority
of such studies examine a limited number of days surrounding the time-to-event (or lag). Thus,
the time-series model considers exposures of short duration, where “short” is defined as a day
or several days or the maximum of a few weeks (Schwartz, 2000).
It should be recognized that deaths associated with pollution events need not be exclusively due
to the effects of exposure to air pollution shortly before death. As suggested in Category I (above)
some of the “air pollution victims” may have already been suffering from a disease enhanced
by past cumulative exposure to air pollution, whereas among other “victims,” the underlying
susceptibility may not have been a consequence of air pollution (Category III) (Quénel et al.,
1999). It should be understood that time-series studies cannot directly model the contribution
of short-term and long-term exposure to air pollution to underlying conditions or frailty of
health status.
The time-series approach has inherent limitations for differentiating between outcomes. If it
were true that death was advanced by only a few days among all “victims,” lifetime lost could
be indirectly addressed from time-series data through assessment of the average time between
the mortality peak and the subsequent rebound (“harvesting”). Despite evidence that harvesting
exists, particularly among those in frail health, it is neither the only explanation nor the most
important explanation for the short-term effects seen in time-series studies (Zeger et al., 1999;
Samet et al., 2000b). The distinction between risk groups experiencing short-term displacement
of mortality (those in frail health) and groups for whom air pollution triggers death much earlier
than would otherwise be expected is difficult to determine, and requires further study (Zeger
et al., 2000). In summary, years of life lost among the pool of “victims” whose deaths cannot
be explained by short-term displacement (harvesting) are not quantified in time-series studies.
4.3
Air pollution attributable death in cohort studies
Cohort studies model the association between an exposure and time to death. The source of
exposure variability is dependent upon the long-term cumulative air pollution across study
emission.
participants. The currently available cohort studies of air pollution (Dockery et al., 1993; Pope
et al., 1995; Abbey et al., 1999) modeled the statistical association between measures of long-term
cumulative exposure and time-to-death. In contrast to time-series studies, the impact of air
pollution on time-to death (survival time) cannot be assessed in one single study area. The
cohort study design permits evaluation of the impacts on health or mortality of repeated
exposure or cumulative long-term exposure and their contribution to changes in time-to-death.
Cohort studies successfully assess the effects of long-term exposure to air pollution because
person-time is the measured outcome the association between exposure and years of life lost
can be derived. Cohort-based effect estimates capture the full amount of time lost across all
three types of air pollution attributable cases (categories I, II, and III), and in distinction to timeseries studies, time lost due to short-term “acute” advancement of death (categories I and III
) cannot be disentangled from time lost due to air pollution-enhanced chronic morbidity
(category II) (Künzili et al., 2001).
4.3.1
Attributable number of deaths
Technically, exposure-response functions from either time-series studies or cohort studies may
be applied to calculate the number of cases that can be attributed to air pollution. In time-series
studies, only two (I and III) of the three categories of air pollution-related cases are captured.
1.
2.
3.
4.4
Death that has been triggered by air pollution shortly before death.
The time-series-based number of attributable deaths, summed across 1 year, indicates
the number of subjects whose deaths have been advanced by air pollution shortly before
death (Samet et al., 2000a). However, it may not be correct to interpret these deaths as
additional cases of death per year. Whether these subjects would have died during the
same year or later, had air pollution levels been lower, requires knowledge about the
average amount of lifetime lost across the air pollution attributable victims. This
information is only supplied by cohort design studies.
Cohort data have very limited power to separately assess short-term effects (categories
I and III).
Mortality and Hospitalization:
There have been a large number of epidemiological studies that have associated particulate
matter with total mortality; with mortality from respiratory causes, and with mortality from
cardiovascular disease. While these have demonstrated considerable variability in terms of
association of potential for exposure to a specific pollutant with death, there can be little
question that statistical analysis and meta-analysis of large population databases has
demonstrated the association of air pollution with mortality. There is a requirement for better
and more sophisticated exposure attribution before the generality of association can be applied
to the individual to describe personal risk. The epidemiology studies have generated many
questions that investigate the biological plausibility for mechanisms that directly associate PM
with biological outcomes and adverse effects. Some of these include changes in heart rate, the
recruitment of neutrophils from the bone marrow referred to elsewhere in this document.
emission.
It has been noted by many investigators that the association between particulate matter and
health impacts does not appear to observe a threshold, or level below which no effects are
observable. Vedal et al. (2003) have examined the associations between air pollutants and
mortality and respiratory disease in Vancouver, B.C., an urban area that has low levels of air
pollution relative to other large urban areas. Deaths resulting from respiratory causes among
Vancouver residents showed a pronounced cyclical pattern when compared to deaths resulting
from cardiovascular causes. Vancouver has a moderate marine climate there are essentially two
‘seasons’, and the pattern of mortality appeared to track this seasonality. Statistically significant
effects on deaths by respiratory causes in the summer were observed for PM10, O3, and SO2 ,
but the effects of NO2 and CO lacked significance. When the results of this study are compared
to other cities with low pollutant concentrations, an increase in air pollutant concentrations are
associated with adverse effects on daily mortality. Although this observation may support
arguments that there are no threshold concentrations below which an adverse effect cannot be
detected, it also brings about the concern that the associations are not reflecting the effect of
the measured pollutants, but rather some factor or combination of factors such as unmeasured
air pollutants or uncontrolled feature of meteorology may be correlated with the measured
pollutants (Vedal et al., 2003).
4.4.1
Recent Epidemiology Showing an Association of Mortality
with Urban Air Quality.
The association of mortality with air pollution events has often identified a particularly sensitive
sub-set of the population that is “harvested” by the conditions. There has been little direct
evidence that all mortality could be attributed to such a select group of individuals. Should the
notion of harvesting be accepted, how many days or weeks or months is death for an individual
advanced by air pollution exposure? In a time-series study in The Netherlands, Brunekreef et
al. (2000a) found a strong association between the day-to-day variation in pollen concentrations
and that of deaths due to cardiovascular disease, chronic obstructive pulmonary disease, and
pneumonia. If there is a harvesting effect from PM10 and its co-pollutants, then the association
between air pollution and mortality should be reflected in a decrease in mortality shortly after
the episode. Also, in the event that harvesting plays an important role, then death rates over the
longer period would show little effect from non-episodic pollution events (Brunekreef et al.,
2000b).
If death is advanced substantially by exposure to PM10 but not necessarily in response to acute
exposure events, then there may be evidence for recruitment into a susceptible population as
a result of a pollution event. For example, a high pollution episode might not result in imminent
mortality, but would be reflected in a steadily increasing death rate in the population for some
period (up to 120 days) after the event. These would be the deaths of individuals weakened by
the first event, or “recruited” a more sensitive population.
Schwartz (2000b) has reported that non-traumatic death from all causes in Boston did not
decrease but increased with the length of time, or window of observation that was considered
for mortality. When subdivided by disease:
emission.
• Admissions for patients suffering from COPD after a pollution event did show clear
evidence of harvesting close to the point of high pollution;
• Deaths from ischemic heart disease appeared to show an accumulative effect of
pollution, represented by an initial rise at the point of high pollutant exposure,
followed by a sustained increase in hospitalizations for some time;
• Mortality due to pneumonia showed some evidence of harvesting, but the data also
suggest additional recruitment over the subsequent weeks. This confirms that
pollution exposure, even for non-critical cases had longer-term impacts that affected
overall longevity.
Thus, after an event of high pollution, the effects on survivors of the initial event is to cause
sufficient damage to increase the risk or possibility of mortality from ischemic heart disease for
a period of up to two months. Those with respiratory illness and COPD are more likely to
succumb to a high pollution event on the day or shortly thereafter, but their risk over the
succeeding time period is little altered by the exposure. The data for this study was drawn from
the early 1980s to mid 1990s, so improving air quality in the intervening period might eliminate
ability to reproduce the observed effects (Schwartz, 2000b)
There is no substantive evidence that would suggest an initiating event (higher pollution day)
recruits healthy people into a pool of susceptible individuals who might be at greater risk from
subsequent pollution event(s) (Schwartz (2000b). Cohort studies (studies that look at selected
populations, and that include accurate data for quantifying exposure) lasting for several years
would be necessary to determine an effect of chronic exposure to pollution for periods of a year
or less. To date there are very few such reports available (Pope et al., 1995b; Künzli et al., 2000).
As the length of time considered after an initial event in a time-series study, the association with
subsequent mortality is subject to confounding, such as influenza or other stressors that give
rise to an increased risk for mortality in a population. Thus the time-series study approach is
inadequate to the task for predicting effects and mortality risks from chronic (or longer term)
exposure to elevated pollution levels.
In the 20 and 90 cities report, Samet et al. (2000b) found an average approximate increase of
0.5% in total non-accidental mortality per 10 :g/m3 increase in PM10 concentration for both
the 20 and 90 cities at 1 day after exposure (1-day lag).
“The PM10 effect was slightly greater for cardiorespiratory mortality than for
total mortality at 1 day after exposure; effects at other times after exposure did
not vary substantially from one another for total or cardiorespiratory mortality.
For both the 20 and 90 cities, the association between PM10 and mortality did
not appear to be sensitive to the inclusion of other pollutants in the model.
Also, when other pollutants (sulfur dioxide, nitrogen dioxide, ozone, carbon
monoxide) were considered for their independent association with mortality in
20 cities, some (CO, NO2) showed associations, but these associations were not
robust to the inclusion of PM10 in the model. The investigators also report that
the effect of PM10 varied across regions in the US, with the largest effect
emission.
observed in the Northeast.”
4.4.2
Morbidity, Hospitalization and Urban Air Quality
Burnett et al. (1999) have described associations for all cardio-respiratory admissions in Toronto
covering a period from 1980 to 1994. There was a strong association between
respiratory/cardiovascular hospital admissions and all pollutants. O3 was linked to respiratory
disease, but not heart disease. CO was linked to cardiac failure. Burnett et al. (1999) have
broken out risks by specific disease as shown below (Table 2).
Table 2: Data for relative increased risk for hospitalization from specific disease
causes (based on time-series) after a doubling of the pollutant level (24-hr
avg). From Burnett et al., 1999
Reason for admission
Association (percent)
Asthma
PM2.5-10 (4%); CO (4%)
COPD
PM2.5-10 (3.86%); CO (3.00%); O3 (6.08%)
Respiratory infection
PM2.5 (6.08%); NO2 (4.44%); O3 (3.93%)
Dysrythmias
PM2.5 (2.47%); CO (7.00%); O3 (3.34%)
Heart failure
CO (4.09%); NO2 (6.89%)
Ischemic heart disease
NO2 (8.34%); SO2 (0.95%)
Samet et al. (2000) found that the concentration of PM10 was positively associated overall with
elderly hospital admissions in the 14 cities for the 3 diseases studied: cardiovascular disease,
COPD, and pneumonia. Although nearly all cities showed a positive association between
admissions and PM10 , the magnitude of the estimated effect on admissions varied considerably
among cities.
On average, cardiovascular admissions increased by about 1% for every 10 :g/m3 of PM10 .
The NMMAPS study examined respiratory illness and obstructive pulmonary disease together,
and showed that pneumonia and COPD admissions rose by about 2% for the same increase in
PM10 (10 :g/m3) (Samet et al., 2000b). The observed effect of PM10 concentration on hospital
admissions persisted in analyses excluding days with PM10 concentrations above 50 :g/m3.
When only days characterized by levels of PM10 lower than 50 :g/m3 were used to derive the
estimates for health effects for each of the 3 diseases (cardiovascular disease, COPD, and
pneumonia) these increased by about 20% by comparison with estimates that included all days
(with higher PM10 concentrations) (Samet et al., 2000b). This suggests that hospital visits on the
day of a high pollution event are less likely to reflect direct effects of the exposure.
4.4.3
Urban Air and Increased Frequency of Admissions for
emission.
Cardiovascular Disease.
The adverse health effects associated with exposure to particulate matter are specific to the
cardio-respiratory tract (Environment Canada, 1999). The cardio-respiratory problems
examined in studies of the effects of PM on human health include lung function, lung
infections, asthma, chronic bronchitis and emphysema (chronic obstructive lung disease) and
heart disease (Environment Canada, 1999).
The health effects measured by increases in hospital admissions for pollutant gases should not
be treated entirely separately, but rather should be combined, because those effects attributable
to PM10 alone will include overlapping domains of effect. This point is most strongly made in
the recent multiple study review carried out by Samet et al. (2000b). Although NMMAPS
focussed on the effects of PM10, examination of the independent effects of other pollutants is
also warranted. Effects on daily mortality were found for most of the gaseous pollutants (SO2,
CO, NO2) in the 20 cities although these effects were generally diminished when the model
controlled for PM10 and other pollutants. In contrast, the association with PM10 did not appear
to be affected by other pollutants (Samet et al., 2000a).
Time-series analysis has been carried out using medical histories of citizens of Buffalo, NY to
assess the effect of acidic PM on daily mortality and morbidity (Gwynn et al., 2000). They found
that the principle association for mortality from all causes was PM10, CO, H+, and SO42-. For
hospitalization risks from all causes, significantly increased risks were found for the following
when present at elevated levels in urban air: H+, NO2, O3 and SO42-. Circulatory mortality was
only significantly associated with PM10. Unlike the Toronto study, there were no statistically
robust associations between single pollutants and circulatory/cardiovascular hospital admissions
(Gwynn et al., 2000).
Schwartz (2000a) has recently compared mortality and pollution data from ten cities in the
United States. It has been argued that deaths occur in a special sub-population of those who
are seriously ill, and would probably have died in a few days anyway. The association between
particulate matter and mortality is stronger and more robust than for any other pollutant.
However, there is accumulating evidence that a public policy that focuses on an action for
reduction of exposures (and therefore the presumed effect of such exposures) on key highpollutant days will have little or no impact on pollutant-related mortality.
The overall correlation between PM10 and mortality according to the study conducted by
Schwartz (2000a) was 0.67% [CI 0.52-0.81] for each rise of 10 :g/m3. This was the same for
summer or winter, but associations for deaths that occurred in hospitals was 0.49% [CI 0.310.68] were lower than the overall non-traumatic mortality that occurred outside hospitals (0.89%
[CI 0.62-1.12]).
Schwartz (2000a) found that for urban air pollutants that occur with PM10 such as CO, SO2 or
O3, there was no statistically significant increase in PM10 levels that could be associated with
increases of one pollutant or another. In other words, no one pollutant gas was a better
emission.
indicator of PM10 compared with any other. Schwartz (2000a) found no reliable observable
effect for increasing ozone/PM10 ratio on the number of deaths that could be calculated using
PM10 alone. Taken together, these results suggest that measuring pollutant gases is not as
important as the measurement of PM10 as a predictor of health effects in general and specifically
for prediction of mortality.
The finding that deaths associated with increases in PM10 occurred more frequently outside
hospitals than inside hospitals in these ten urban areas suggests the sensitive population is not
limited to those already close to death. Thus sudden (cardiac and other) death appears to be a
major component of air pollution-associated risk (Schwartz, 2000a). This is accordance with
a recent report by Liao et al. (1999) who found that fine PM2.5 can disturb heart rhythms by
increasing arrhythmia in cardiac patients, and for similar effects induced in dogs by application
of suitable experimental models (Godleski et al., 2000).
Schwartz et al. (2002) have argued that there is no evidence of a threshold for mortality
associated with exposure to PM 2.5 derived from mobile sources. At lower concentrations, the
PM10 slope for association with mortality increases, indicative of a non-linear response from
higher to lower doses (Schwartz, 2000a). This steep slope for increased health effects at lower
doses suggests that there may be a threshold effect, but probably at doses much lower than 50
:g/m3 for PM10. This also suggests that intervention strategies that lower the average levels of
PM10-related pollutants rather than those activities that focus on peak pollution periods may
have the greatest health benefit for the population.
Recently, Dominici et al. (2002) have revised assumptions used in the generalized additive
models (GAM) method for analysis of data from the National Morbidity, Mortality, and Air
Pollution Study (NMMAPS). The GAM method is a flexible and effective technique in wide
use for conducting nonlinear regression analysis in time-series studies of the health effects of
air pollution. They determined that through an error in the default settings of the statistical
software, a bias was introduced into the data analysis. They concluded there was evidence that
the rate of mortality estimated in metanalysis of data from ninety cities (Samet et al., 2000b)
could be approximately half of the original estimated rate of 0.4% increased mortality per
incremental increase of 10 :g/m3 of PM10 (i.e 0.2% based on a one day lag). Although the
reanalysis showed a slope reduction in the linear regression concentration response curve, the
authors assert that there is no evidence to support a threshold effect, even at very low exposures
(Dominici et al., 2002).
emission.
5.0
CHRONIC EXPOSURE TO PM10 AND RESPIRATORY
HEALTH
5.1
Introduction
Current health effects research has not been able to determine a concentration threshold below
which PM does not affect cardio-respiratory health (Environment Canada, 1999; MOE, 1999).
Health effects have been observed at very low levels of PM, and these effects appear to increase
steadily as the sulfate content of particle concentrations rise. On the basis of this result, Health
Canada has concluded that there is no safe level of PM that is not associated with an adverse
effect.
Decreased lung function and increased respiratory symptoms are associated with increased PM
concentrations in community epidemiological studies and controlled laboratory exposure studies
of laboratory animals and humans. Particularly noteworthy is the observation of these
PM-associated effects in children.
There has been a general recognition that chronic exposure to urban air pollution has a
potentially deleterious effect on children’s respiratory health. While epidemiology studies based
on hospitalization or clinical signs of respiratory illness provide suggestive evidence for the
relationship, more direct evidence of reduced lung function in large groups of affected children
was lacking. Recently, a number of studies have begun to appear in the open literature that
specifically address the relationship between particulate matter and reduced lung function
and/or changes in other biological endpoints indicative of links between environment and
respiratory health. Some of these studies are discussed below, and summarized in Table 5
(below). Overall the results of the studies of respiratory symptoms are not entirely consistent.
Nevertheless, there is growing evidence for an association between reductions in lung function
and chronic exposure to urban air pollutants including PM10 using defined populations of
subjects for study.
Künzli et al. (2001) have estimated that a long-term (15 or more years) difference in exposure
to PM10 based on an incremental change of 10 :g/m3 is associated with a relative risk (RR) of
dying (for natural causes among adults aged 30 years or older) was 1.043.
RR =
1 + 0.043 x excess pollutioninincrements of 10µ g / m3
Crosignani et al. (2002) have used this relationship to calculate the excess deaths for long-term
exposure to air pollution. The “attributable fraction” will be (RR-1)/RR assuming residents of
an area are equally exposed. For example, the RR of death for someone living long term in
Hamilton (25 :g/m3 annual average PM10 in 2000 at STA # 29300) instead of another city with
an annual average PM10 of 28 :g/m3 would be calculated as 1 +(28-25)/10 x 0.043 = 1.013.
The corresponding attributable fraction of the mortality is (1.013 - 1)/1.013 = 1.3 % (Crosignani
et al., 2002). That is to say in two cities of equal population where the difference in the annual
emission.
average 24-hr exposure to PM10 was 3 :g/m3, the rates of mortality would differ by 1.3% when
examined over many years, and this would be attributable to pollution. It is important to
recognize that for this generalization to apply, such differences would involve large urban areas,
and large populations, not to localized areas of population.
For purposes of demonstration, we can assume deaths from natural causes in Hamilton are
comparable to the New City of Toronto (GTA). Burnett et al. (1998) reported that between
1980 and 1991, the daily mortality in Hamilton was 2.02 per 100,000, and the Hamilton census
area had a population during this time of 420,000. Recent Statistics Canada data (July, 2002)
shows the total population for the City of Hamilton at 490,270. If we assume that the mean
number of deaths per day from all causes in Hamilton has remained approximately the same,
then the daily mortality based on the new census data would be 9.9 persons per day or 3,622 per
year. In the Toronto GTA, approximately 45% of daily mortality from all causes is related to
cardiovascular or respiratory causes (20.5 deaths of 45.6 deaths per day). On the basis of a
comparison with Toronto, of the 3,622 deaths per year in Hamilton from all causes 1,652
(45.1%) could be attributed cardiovascular and/or respiratory causes.
On the basis of this association between level of pollution and mortality, were the annual
average 24-hr concentration of PM10 in Hamilton to change by 3 :g/m3 to 28 :g/m3 the
expected increase in total mortality from all causes that would be attributable to air pollution
would be 1.3%. Similarly, a decrease on annual 24-hr PM10 of this magnitude would reduce all
cause non-traumatic mortality by a similar amount. Four important facts about this relationship
need to be recognized:
• This relationship is based on exposures of many years, so the gains in improved health
and reduction in mortality will occur slowly, and not be immediately noticeable.
• As air quality improves and approaches background (about 15 :g/m3 in Canada) it
becomes increasingly difficult to attribute or to distinguish the causes of mortality.
• Mortality is not solely the result of the cumulative impact of environmental pollution.
• These relationships are based on very large populations and on averages, so they are
poor predictors of individual risk from local fluctuations in pollution levels.
A 1.3 % increase in deaths from non-traumatic causes in Hamilton using this formula and an
increase of 3 :g/m3 in overall annual PM10 would equal 47 persons on account of the increased
level of particulate matter. If we attribute the impacts of PM10 largely to causes of respiratory
and cardiovascular death, then the difference in the death rate from these causes would be 1.3
% of 1,652 deaths, or 21.5 deaths per year. In other words annually and for each 22,800
residents of Hamilton the death of one person a year attributed to cardiovascular or respiratory
causes could be said to result from air pollution PM10.
Thus, an overall improvement in air quality of 3 :g/m3 over all of Hamilton could be said result
in a appreciable change in the quality of life. As pointed out above, it is important to recognize
that this difference might be difficult to observe because of the variation in the annual death
statistics, and over time, a reduced mortality might not be immediately apparent. Of course, the
actual adjustment in mortality would not be as evident as indicated in the calculation because
emission.
such a dramatic change in regional air quality would take some time to accomplish.
5.2
Health Impacts from Long-term Exposure to Air Pollution
Long-term impacts of air pollution on human health have relied on the results of cohort
epidemiological studies to evaluate chronic respiratory effects (Zemp et al., 1999), reductions in
lung function development (Gauderman et al., 2002; Horak et al., 2002), mortality in nonsmokers (Abbey et al., 1999), and lung cancer (Reynolds et al., 2002; Pope et al., 2002).
5.2.1
Respirable Fine Particulate Matter PM2.5
Although most of the recent epidemiological research has focussed on effects resulting from
short-term exposure to polluted environments, several studies of longer-term exposures suggest
significant potential for raising public health concern (Pope et al., 2002). The American Cancer
Society (ACS) study was one of two original prospective cohort study populations whose
assessment was initiated in the 1970s. In the most recent review of the cohort, PM2.5 as well as
PM10 was examined for health impact in three categories of mortality: (1) cardiopulmonary, (2)
lung cancer, and (3) all other causes.
Fine particulate matter air pollution has declined over the period of the study in the 51 cities for
which data were available. Cardiopulmonary and lung cancer mortality showed associations with
fine particulate matter (PM2.5) in the ACS cohort (Table 3). “All other cause” mortality did not
show a significant association with this air pollutant (Pope et al., 2002). Weaker less consistent
associations with mortality were observed with PM10 and PM15. PM10 and total suspended
particulate (TSP) and gaseous pollutants were generally not associated with increased mortality
(Pope et al., 2002). Each 10 :g/m3 elevation in long-term average PM2.5 was associated with
approximately 4% (all cause), 6% (cardiopulmonary) and 8% (lung cancer) mortality. The
updated cohort covered a period of 16 years, tripling the number of deaths initially reported in
1995, but the associations between particulate matter and terminal disease causing death
remained. After controlling for smoking behaviour, both cardiopulmonary and lung cancer
mortality remained a significant cause of death for non-smokers.
emission.
Table 3:
Relative Risk (RR) for Mortality Associated with 10 :g/m3
Change in PM2.51
Adjusted RR 1979-19832
[95% Confidence Interval]
Adjusted RR for
Smokers3
[95% Confidence Interval]
All-Cause
1.04 [1.01-1.08]
2.58
Cardiopulmonary
1.06 [1.02-1.10]
2.89
Lung Cancer
1.08 [1.01-1.16]
14.8
All other Causes
1.01 [0.97-1.05]
-
Cause of Mortality
1
From Pope et al., 2002
Estimated and adjusted based on the baseline random-effects Cox proportional hazards model, controlling
for age, sex, race, smoking, education, marital status, body mass, alcohol consumption, occupational
exposure and diet.
3
Average current smoker (men and women combined) 22 cigarettes/day for 33.5 years, with initiation
before age 18 years
2
The shape of the dose response function for mortality with PM2.5 appeared to be nearly linear.
The risk imposed by exposure to fine particulate air pollution is obviously much smaller than
the risk associated with cigarette smoking as shown in column three of Table 3 (Pope et al.,
2002). Pope et al. conclude that long-term exposure to particulate air pollution in metropolitan
urban communities across the United States is an important risk factor for cardiopulmonary
mortality. In addition, the large cohort and extended follow-up of nearly 1.2 million individuals
in 228 communities suggests that fine PM2.5 is responsible for increases in lung cancer incidence
after prolonged exposure (Pope et al., 2002).
5.2.2
Inhalable Particulate Matter PM10 and Criteria Pollutants
Abbey et al., (1999) also updated the Seventh Day Adventist study cohort initially described in
1991. This cohort was of 6,338 nonsmoking individuals originally enrolled in 1977. The
primary pollutants of concern were PM10 (rather than PM2.5) and ozone. Among this group,
long-term exposure to increased level of PM10 (43 days per year greater than 100 :g/m3) was
associated with mortality that included non-malignant respiratory disease (Table 4). Strong
associations were also observed between PM10 and lung cancer in males (Abbey et al., 1999).
Among females, none of the pollutants examined showed a positive association with natural
cause mortality. Unlike the results recently reported by Pope et al. (2002) that showed significant
associations with cardiopulmonary mortality, this Abbey et al. (1999) reported no significant
associations for this endpoint for either sex. Furthermore, the associations in some cases gained
significance based on increases in interquartile ranges (e.g 100 :g/m3 PM10). The association
with PM10 for non-malignant cause for respiratory disease among the Seventh Day Adventists
group was approximately similar to the risk for mortality among former smokers and never
smokers. Exposure to ozone (550 hours/yr above 100 ppb) was also associated with increased
emission.
risk of lung cancer, but only in males.
Table 4: Adjusted Mortality Relative Risks by Cause of Death for an
Interquartile Range (IQR) Difference in Specific Pollutants.
PM10 above 100
:g/m3
IQR
PM10 Mean Conc.
O3 above 100
ppb
24.08 :g/m3
551.1 hours/yr
2.38 [1.42-3.97]##
3.36 [1.57-7.19]^^
4.19 [1.81-9.69]##
1.08 [0.55-2.13]
1.33 [0.6-2.96]
1.39 [0.52-3.67]
43 day/yr
Cause of Death
Non malignant respiratory
disease (Males)
1.28 [1.03-1.57]**
Non malignant respiratory
disease (Females)
1.01 [0.91-1.33]
Lung Cancer (Males)
Lung Cancer (Females)
Neither PM10 nor ozone showed a statistically significant association for health impact in women
(Abbey et al. 1999)
** p <0.05
^^
p < 0.01
##
p < 0.001
In another recent study of long-term effects of ambient air pollution in Switzerland, Zemp et
al. (1999) reported positive associations between annual mean concentrations of NO2, and PM10
for reported prevalence of chronic cough and phlegm production. They also observed
associations for breathlessness while at rest (day or night) and dyspnea (difficult or laboured
breathing) on exertion. Positive associations between increased average annual PM10 (10 :g/m3)
and risk of respiratory symptoms of cough with the production of phlegm was 1.27 [1.08-1.50].
The predicted effect of a 10 :g/m3 increase in annual mean concentrations of PM10 was
substantial. An increase of 35% in prevalence of chronic phlegm; an increase of 50% for
breathlessness during the day and of 33% for breathlessness during the night, and finally an
increased prevalence of 32% for exhibiting dyspnea on exertion (Zemp et al., 1999). Other
symptoms of asthma (wheezing) were not affected by increased PM10. In this study, Zemp et
al. found no evidence for an association of the prevalence of physician diagnosed asthma or for
key symptoms of asthma with the level of long-term air pollution experienced in the Swiss study.
This could well be because the annual mean and the ranges of pollutants observed were below
those commonly observed in many urban areas including Hamilton.
Smaller effects on all of these endpoints were observed with increased annual exposure to NOx
of 5ppb (10 :g/m3). Since the production of oxides of nitrogen occur with the combustion of
fossil fuels, it is possible that the investigators monitored effects of mobile sources of emission.
emission.
Nevertheless, these investigators concluded that long-term exposure to air pollution (PM10 and
NOx) even at low level is associated with higher prevalence of respiratory symptoms, especially
breathlessness, and a symptom of combinations suggestive or related to chronic bronchitis.
These health impacts were strongly affected by smoking and environmental tobacco smoke.
5.2.3
What endpoints describe respiratory health and chronic
exposure of children and adults to particulate matter?
Table 5 reviews some recent publications that compare PM10 with evidence from assessment
of specific toxicological endpoints and the respiratory health in children. In each of these
studies, the authors selected populations that had lived either in areas with greater opportunity
for PM10 exposure (urban) or in cleaner (rural) environments. The combined ambient
monitoring data with respiratory lung function spirometery data such as peak expiratory flow
(PEF), forced vital capacity (FVC) and forced expiratory volume in one second (FEV1). Other
measurements included examination of bronchial hyperresponsiveness (BHR) and
immunoglobulin E (IgE) to determine respiratory health of populations with asthma and
respiratory allergies in children and cardiac obstructive pulmonary disease (COPD) in adults.
emission.
Table 5: Health Effects Associated with Chronic Exposure to Particulate Matter among Children and
Adults
Population & Number
Conditions of Urban Exposure
Endpoints Examined
Results
Reference
CHILDREN
Preadolescents
Age 9-11 years
Total subjects:1001
Exposed: 52.6 ±
53.98 :g/m3
SPM
Control: 33.23 ±
35.99 :g/m3
SPM
FVC, FEV1
measure rate of development for
lung function based on changes
over a 2 yr exposure
Asthmatics were excluded;
Two populations from higher or lower
polluted areas.
Boys & girls from higher pollution areas
showed lower adjusted mean values of
lung function growth.
Only effects on boys were significant.
Jedrychow
ski et el.
(1999)
Kraków, Poland
Rural:
24 h
avg
PM10/m3 (range)
44.7 (4.8-104)
44.1 (7.9-242)
26.6 (7.1-96.6)
PEF (peak expiratory flow)
FVC and FEV1
BHR (bronchial
hyperresponsiveness)
Serum IgE for allergies.
Compare lung function,
prevalence of upper and lower
respiratory symptoms of children
with BHR and effect of pollution
Children with BHR and high IgE had
increased numbers or reported respiratory
symptoms. These children should be
targeted for health initiative.
About 26% of children in the study
population were high for both IgE and
BHR.
Boezen et
al. (1999)
Children 5-14 years
2470 total subjects
Eastern Germany
Polluted
Industrial Urban:
(24h
)
40 :g PM10/m3
Non-polluted
Urban: (24h
)
33 :g PM10/m3
PEF
FVC and FEV1 (total of 1028
children)
Allergen response (IgE)
Respiratory symptom
questionnaire
No differences for any lung function test.
Children from industrial areas were more
sensitive to allergens.
Industrial cities had increased incidence
of asthma (2.1 and 4.4% increase)
Clean city had 1.6% asthma frequency.
Industrial cities had higher values for:
Wheezing OR = 1.79
Shortness of breath OR = 2.36
Coughing/cold OR = 1.72
Heinrich et
al. (1999)
Children
Total Subjects: 3676
50% 9-10yr
25% 12-13yr
25% 15-16yr
34.8 :g/m3
Range 13-70.7
:g/m3 (24h
)
Children selected from 12
Southern California
communities evenly distributed.
Questionnaire; cough wheeze,
asthma.
No toxicological endpoints
Children with Asthma responded to
increases in PM10 with increased
symptoms.
Same for increased NO2.
McConnell
et al.
(1999)
0
Urban: 3
winters,
PM10/m3 (range)
58.4 (4.7-145)
41.5 (12.1-113)
31.1(8.8-90)
0
0
Children: 7-11 Years
459 subjects
Netherlands
0
emission.
Adults
Population & Number
Endpoints Examined
Results
Reference
Compare location of residence to
busy roads in SanDiego, Calif.
Traffic counts and hospital visits.
No significant association between any
health endpoint and traffic. 75% lived
within 550 ft of a busy road.
No evidence of increased asthma risk
with higher traffic counts. Among those
diagnosed as asthmatic near a busy
street was responsible for small increase
in hospital visit. May have been better
health management and not pollution.
English et
al. (1999)
10-33 :g/m3
PM10. Average
of 5 years
residence.
Switzerland
Cough, Bronchitis, nocturnal
cough symptoms monitored
(Diary)
Questionnaire linked to ambient
monitoring data (daily).
Comparison of frequency of
symptoms with monitoring data
for children in different
communities.
Significant association between annual
mean levels of PM10, NO2, SO2 and
respiratory symptoms.
Long-term exposure probably not a
causal factor for asthma or development
of respiratory allergies.
BraunFahrländer
et al.
(1997)
Children
3-15 yrs (71%)
<2 yrs (29%)
Average 565 visits/day
over 6 month period
Santiago, Chile
Urban exposure
(24h
)
108.6 :g
PM10/m3 range
18.5-380
Total number of clinic visits for
respiratory symptoms. Upper
and Lower respiratory
symptoms; All data based on
clinical appraisal
Under two years, an increase of 50 :g
PM10/m3 resulted in a 4-12 % increase in
clinic visits for lower respiratory;
for 3-15 yr olds, an increase of 50 :g
PM10/m3 resulted in 3-9% increase in
lower respiratory symptoms.
Ostro et
al., (1999a)
Children
89 asmatics over 7
months
47 :g/m3 (range
3-171)
SO4, strong particle acidity, SO2,
Cough, dypsonea, phlegm,
runny nose. PEF morning &
evening
5-day mean increase in SO4 of 5 :g/m3
produced a 0.48% decrease in PEF. For
all pollutants *same day) 0.06% decrease
in PEF. Evening PEF effected.
Peters et
al., 1999a
5996 < 14 years
diagnosed
asthmatics
Children: 6-15 yrs
4470 subjects
0
Children (Asthmatics
and respiratory
symptom free)
8280 subjects
2284 diagnosed
with nonrespiratory
visits for
medical
attention
0
emission.
Adults
Population & Number
Children
8-11 years
59 measurement days
over 11 months
S.W. Mexico City
Urban 24h
average
PM10 49 :g/m3
(14-87 :g/m3
range)
PM2.5 30 :g/m3
(9-69 :g/m3
range)
Endpoints Examined
PEF, PM, and O3 before and
after school day.
emission.
Results
Reference
Exposure predicted phlegm. There was a
combined effect of O3 & PM. O3 has an
immediate effect on lung function. PM
effects lagged several days. Chronic
exposure to O3 in previous 1-2 wk affects
lung function. Combined effect is 7.1%
reduction in PEF (3.9-11.0%) after 7 days
exposure PM2.5 (17 :g/m3), O3 (25 ppb)
Gold et al.,
1999
Adults
Population & Number
Endpoints Examined
Results
Reference
ADULTS
Suburban (24h
) (Marseilles) 3
62 :g PM10/m
range 93 :g
PM10
101 avg age 52
y
)
13 :g PM10/m3
range to 50 :g
PM10/m3
168 avg age 56
y
0
Urban:
Amsterdam, NL
24 h
44.1 :g PM10/m3
(12.1-112.7)
91 subjects
37% male
Rural:
Meppel, NL
41.5 :g
PM10/m3
(7.9-242)
98 subjects
50% male
0
Adults: All with airway
lability (BHR or PEF
variability)
Total 189 avg age 60
yrs1
Downtown (24h
0
Adults: 269 subjects
ONLY COPD or Asthma
patients
10 or more years single
residence1
PEF
FVC and FEV1
BHR (hyperresponsiveness)
reversibility of obstruction with
$2 agonist (salbutamol)
Questionnaire: Asthma, Rhinitis
Interview- test lung function, $2 agonist,
retest to confirm BHR patient.
PEF
FEV1
Self reporting diary:
Cough, Phlegm, Upper
respiratory, Lower respiratory
symptoms
Pollution not significantly different, but
elevated in urban area
For non-compromised adults there was
NO consistent positive association
between prevalence of respiratory
symptoms and increases in levels of air
pollution.
For adults with BHR or variable PEF,
there was an INCREASED prevalence of
respiratory symptoms among those living
in an urban area.
emission.
Jammes et
al. (1998)
BHR/Asthma response greater downtown.
Predominantly males affected females
were same up or down town. Responses
in males accounted for all the significant
differences.
All subjects were chosen for BHR
(Asthma or COPD). Most strongly
affected by PM were urban males.
Boezen et
al. (1998)
Adults
Population & Number
24 h
21.2 :g PM10/m3 for all sites
Range 10.1 to 33.4
Adults
10 year prospective
study in California
(Santa Clara County)
PM10 %54.1 :g/m3
& 52.7 :g/m3
Days PM10 >100 :g/m3:
% 31.3 days
& 35.3 days
0
Adults 18-60 yrs
8 cities (Switzerland)
9651 subjects
Endpoints Examined
Results
Reference
FVC and FEV1
IgE (allergy)
Questionnaire: Wheeze, Asthma,
Chronic Bronchitis, smoking
3-4% change in FEV1 per 10 :g PM10/m3
increase in exposure.
Population figures segregated by location
of residence.
The towns investigated, and their
populations were compared based on
pollutant levels.
A majority of the resulting correlations
with polluted cities was due to Geneva
and Lugano, the only locations with
appreciable PM10 over 10 :g/m3 ambient
concentration.2
AckermanLiebrich et
al. (1997)
For 54 days per year when PM10 exceeds
100 :g/m3 there is a 7.2% reduction in
FEV1
For SO4 increase of 1.6 :g/m3, there is a
decrease in FEV1 of 1.5%, but only in
males.
There is a strong association between
affected individuals and history of
respiratory disease in the family.
Abbey et
al., 1998
1
The Jammes and Boezen studies on adults both show that for compromised adults, (those with pre-existing disease identified as asthma or BHR) biological
responses are greater for those living in urban areas compared to those from rural areas. For adults with no preexisting disease, the incidence of respiratory
symptoms does not greatly differ between the urban and rural locations. BHR affected males may be more sensitive to pollution and PM than women.
2
In spite of an environment low in PM10, it was possible to identify pollutant related effects for lung function in exposed populations. IT IS IMPORTANT TO
NOTE that the % difference for FVC from expected was unchanged for all comparisons, or was POSITIVE up to about 20 to 25 :g PM10/m3. Only reductions
in these lung function tests were noted in Lugano and Geneva. This is a possible threshold, since effects with a negative impact for health were observed
in the larger cities.
emission.
Jedrychowski et el. (1999) examined over one thousand children in urban and rural areas near
Kraków, Poland for effects of chronic exposure to suspended particulate matter on the
respiratory health of a cohort of boys and girls. There was a significant difference in the annual
mean 24-hr value for particulate matter between the urban (52.6 :g/m3) and the rural (33.3
:g/m3) environments of nearly 20 :g/m3. Lung function measurements were taken two years
apart. Children were recruited at age 9, and reexamined at age 11. Adjustments were made for
the rate of growth and maturation of the lungs and ancillary respiratory tissues. Children
specifically diagnosed as asthmatic at the beginning of the study were excluded. After two years,
both boys and girls living in the urban polluted area had reduced lung function growth, but only
the urban boys showed a statistically significant reduction.
Boezen et al. (1999) in the Netherlands also examined respiratory function of children over three
winter periods to determine whether there was any effect of urban versus rural exposures. The
subjects ranged in age from 7-11 yrs, and they were examined for bronchial hyperresponsiveness
(BHR), standard lung function parameters and serum IgE for allergies. Measurements of lung
function (PEF) were recorded daily for a period of three months along with upper and lower
respiratory symptoms. The most significant response to elevated PM10 (black smoke) and other
air pollutants was among children who demonstrated reactive airways (BHR) and elevated
immunoglobulin E. This group represented 26% of the 459 subjects. There was no clear
difference between responding children from urban and rural areas, but the difference in the 24
h mean values for PM there was a significant increase in lower respiratory symptoms for every
100 :g/m3 in children with both elevated IgE and positive for BHR. No other groups of
children (classified as BHR+, and or IgE+) produced data to support decreased respiratory
function from pollutant exposure. Thus those with pre-existing asthma or allergy related
respiratory problems could be at risk and therefore introduced to a health management program.
Heinrich et al. (1999) examined 2470 subjects aged 5-14 years who lived in one of three towns
located in what was previously East Germany. Two towns were industrialized, while a third was
considered not impacted by industrial pollution. While these towns did not differ greatly as to
PM10 (24 h mean) (Table 5), a series lung function tests were carried out to determine whether
there were notable differences in response to prolonged elevated exposures of PM at any of the
sites. There were no differences of significance among the test populations for respiratory lung
function. Children from industrialized populations showed an increased immunoglobulin level,
and about twice incidence of asthma (1.6% for the clean city, 2.1 and 4.6% for the industrial
cities). Members of the study group from the industrial cities had increased frequency of
reporting lower respiratory symptoms. Thus, although there were significant health effects that
arose as a consequence of growing up in an industrialized town, it is uncertain as to the likely
cause, and it was probably not high levels of PM.
Peters et al. (1997b) have examined the effect of PM on respiratory morbidity in 89 asthmatic
children in central Europe (Czech Republic). Asthmatic children from Sokolov, CR showed a
small, but statistically significant association between lung function and pollution in the winter
months. Respirable particulate (measured as SO42-) was associated with a reduction in PEF.
emission.
SO2, and strong particle acidity, on the other hand, also showed significant but weaker
associations. The combined effect of PM and other pollutants resulted in a decrease if 0.06%
in PEF for a 10 :g/m3 increase in PM10. However, an increase in the five day mean
concentration of SO42- by 5 :g/m3 was associated with a 0.48% decrease in lung function
parameters. Therefore, prolonged exposure to sulphates was statistically associated with
decreased lung function.
In another study of 2564 school aged children (7-12 yrs) in Finland, Pekkanen et al. (1997)
concluded that it was only necessary to follow on asthmatic children enrolled in a study of
chronic respiratory effects (cough, bronchitis, dyspnea, etc) associated with the exposure in an
urban environment. They reported on 39 of all the children (18 girls and 21 boys) were selected
for follow-up after filling out a questionnaire. These individuals were followed for a two month
period in the winter months during which simultaneous meteorological and pollutant
measurements were collected. Students received lung function tests and were assessed during
this period. Twenty-one children were diagnosed as asthmatic, and 30 were identified as atopic
(positive allergen response).PM10, NO, NO2 and CO were monitored, and it was concluded that
daily variation in PM was correlated with vehicle traffic. There was high intercorrelation of fine
PM (#0.32 :m), NO, NO2, and CO. However, the reductions in PEF observed in the children
correlated equally well with particles of several size ranges, and no differences were noted that
could distinguish PM10 from any size group of PM examined. Therefore, it was concluded that
in an urban setting where PM was primarily due to local vehicle emissions, there was no
correlation of sufficient strength with any other source os combustion emissions. Since the
study took place in the winter months, in Northern latitudes, there was no likelihood for
production of significant productions of ozone from anthropogenic sources.
Abbey et al. (1998) reported on a prospective study involving 1391 non-smokers with 20 years
of exposure to ambient concentrations of respirable particles, suspended sulphates, sulphur
dioxide, ozone and indoor air particles in Santa Clara County of California. Changes were noted
in lung function, but these changes were not attributable to short term exposure to air
pollutants. They found that the parental history of respiratory disease was associated with an
increased susceptibility to the effects of air pollution in males.
Abbey et al. (1998) found decreased FEV1 in this group was associated with PM10. An increase
in the number of days/year when PM10 exceeded 100 :g/m3 by 54 days was associated with a
decrease in FEV1 of 7.2%. This change in lung function, however, was only associated with
males whose parents has asthma, bronchitis, emphysema or hay fever. SO42- was also associated
with decrements in FEV1 for all males whatever their parental history. A 1.6 :g/m3 increase in
sulphate was associated with a 1.6% decrement in FEV1. Ozone (23 ppb increase) as an 8h
average was associated with a 6.3% decrement in FEV1, but again only in males who had a
familial link to respiratory problems.
Peters et al.(1999a,b) have described a 10 year prospective study of school age children in
Southern California. Pollutants assessed includes ozone, acids, NO2 and PM. Twelve
emission.
communities were involved in the study, with 150 grade 4, 75 grade 7 and 75 grade 10 students
from each of the community study areas. Risk factors identified included a greater frequency
of respiratory illness in males, risks fro associations with pets, mildew and water damage,
associations with parents who has a history of asthma, and elevated risks from association with
environmental tobacco smoke. In boys only, there was a positive association for increased
prevalence for wheeze due to acid particulate (OR=1.45 [CI 1.14-1.83]) and to exposure to NO2
(OR=1.54 [CI 1.08-2.19]). After adjustment PM10, PM2.5, and NO2 were each associated with
lower FEV1 and other measures of degraded lung function performance. Acid vapour in urban
air was associated with decreased FVC and FEV1, while ozone was associated with lowered peak
expiratory flow rate (Peters et al., 1999b).
Lee and Shy (1999) examined six communities in the Southern USA for evidence of increased
risk of lung function decrements due to three types of waste incinerators. Populations exposed
to emissions from each type of incinerator (biomedical, municipal and hazardous waste) were
matched with unexposed populations. The objective of the community-based diary study was
to determine whether respiratory effects of air pollution could be identified in any of the
communities. They wished to determine if PM10 exposure is associated with respiratory
function as measured by peak expiratory flow rate (PEFR); whether indoor air sources were
associated with such effects; and could any effects on lung function be attributed to the presence
of an incinerator in the community.
Persons recruited for the incinerator exposure study, and who were classified as respiratory
hyperresponsives showed a slightly lower PEFR on average than normal subjects. While the
study did show that hyperresponsive populations may show increased susceptibility or variability
to air pollutants, there was no specific association with the location of an incinerator in their
community. The study clearly shows that hyperresponsive individuals differ in degree of
response or air pollutants in general. There appeared to be greater observable effect from
regional air pollution exposure (long range transport) than exposure to incinerator sources (Lee
and Shy, 1999).
The association of ozone and PM on the lung function of school aged children has been
examined in Southwest Mexico City (Gold et al., 1999). The investigators assessed the effect of
PM10, PM2.5, and ozone on the lung function of 40 school children (8-11 years). Each child was
administered a lung function test in the morning before school commenced , and in the
afternoon before returning home. This was repeated at intervals (59 times) over more than 11
months, so all seasons could be examined. SO2 levels in the community were low because of
the low level of industrial activity. NO2 data was insufficient. The results of the peak flow lung
function tests show an immediate effect of ozone exposure. The effect of elevated PM was not
observed until sometime (lag time) between exposure and effect. This lag was most notable on
performance during exercise. When ozone was chronically elevated, increased exposure to PM
as well as ozone during the previous 1-2 weeks can have a demonstrable effect on lung function.
The cumulative effect of PM2.5 and ozone exposure in Southwest Mexico City beyond the
interquartile range (17 :g/m3 PM2.5, and 25 ppb O3) resulted in a reduction of PEF of 7.1% [CI
emission.
3.9-11.0] (Gold et al., 1999).
The effect of ozone and PM has been examined in other communities of school children. Chen
et al. (1999) examined 941 school aged children for effects of urban air pollutants on lung
function. Measurement endpoints included FEV1, FEV25-75, PEF, etc. These were matched
against the independent variables of PM10, CO, SO2, O3 and NO2. Children were selected from
rural, an urban/light industrial neighbourhood, and from a community living near a
petrochemical/refinery facility. The peak ozone concentration on the previous day was found
to be an important factor responsible for any lung function decrement in FEV1, or FVC. It was
not possible to attribute any reduction in lung function parameters to a pollutant source, leading
the authors to conclude their measurement tools were of insufficient sensitivity.
5.2.4
What are the associations between other clinical effects in
children and chronic exposure to urban air based on PM?
Several recent reports specifically examine groups of children exposed to urban air pollution to
develop a relationship between chronic exposure as defined by PM10 and respiratory symptoms.
In a large study, McConnell et al. (1999) selected children of different age groups across twelve
communities in Southern California, and compared the frequency of respiratory symptoms with
indicators of ambient pollution (Table 5). The mean 24h exposure for the more than 3600
children in the study was 34.8 :g/m3. They found that there was evidence of a consistent
relationship between PM10 and increased respiratory symptoms among diagnosed asthmatics.
There was no significant difference for children without pre-existing respiratory conditions.
Braun-Fahrländer et al. (1997) examined 4470 children aged 6-15 years who lived for an average
of five years in one of ten communities in Switzerland. The ambient levels of pollutants (24 h
mean for PM10 of 10 to 33 :g/m3) were very low compared with other locations Europe,
particularly in the East. The strongest relationship with adverse respiratory symptoms was
found for PM10, but only in conjunction with other pollutants. It was not possible to identify
or associate health risk with specific components of an environment of mixed air pollutants.
The authors did conclude that there was no evidence that long term exposure to PM10 was a
causal factor in the development of allergies or asthma. In another recent study involving
clinical assessments of children in Santiago, Chile during the winter months, an association was
found between increased PM and respiratory symptoms (Ostro et al., 1999a). Santiago is a city
that suffers from acute pollution problems, where the mean 24 h PM10 during the winter months
reaches 108 :g/m3 (Table 5). A rise of 50 :g/m3 in the PM10 was associated with a 4 to 12%
increase in the number of lower respiratory symptoms of children under two years, and a rise
of 3-9% in the frequency of symptoms for children 3-15 yrs old. There can be no question that
chronic exposure to PM in this range has a demonstrable adverse effect on the respiratory health
of young children.
In summary, there is mixed evidence for increased incidence of respiratory symptoms from
PM10 exposures among children not previously diagnosed with asthma or allergies. Where
ambient levels are under 50 :g/m3, there is conflicting evidence for increased respiratory
emission.
symptoms with chronic exposure. The California study (McConnell et al., 1999) may not be
representative of what might be expected in Eastern Canada or the U.S. North East where
sulphate contributes more to total PM mass. The Ontario Ministry for the Environment (MOE,
1999) has supported the position of Health Canada outlined in the Scientific Assessment
Document for PM (Environment Canada, 1999) that “population studies based on large
administrative databases (such as the hospital admissions study in Southern Ontario based on
a population of 8.7 million people) are able to demonstrate impacts of pollution on public
health.”
5.2.5
What Are the Effects of PM10 on the Lung Function and
Incidence of Respiratory Symptoms in Adults?
Recent studies that assess the toxicological effect of increased PM10 in adults has focussed on
those with pre-existing disease. Jammes et al. (1998) and Boezen et al. (1999) (Table 5) studies
on adults both show that for compromised adults, (those with pre-existing disease identified as
asthma or BHR) greater biological responses are observed for those living in urban areas
compared to those from rural areas. For adults with no preexisting disease, the incidence of
respiratory symptoms does not greatly differ between the urban and rural locations. BHR
affected males may be more sensitive to pollution and PM than women. This echos the
observation of Jedrychowski et al. (1999) who identified young males as most significantly
affected by urban pollutants including PM10. Ackerman-Liebrich et al. (1997) examined a group
of 9651 subjects living at eight different locations across Switzerland; a country with low levels
of ambient pollution. In spite of an environment low in PM10, it was possible to identify
pollutant related effects for lung function in exposed populations. It is important to note that
the outcomes of lung function tests did not show a pattern of linear dose response reduction
in FVC coordinated with increased PM10. Expected values for lung function parameters
remained essentially unchanged for communities up to about 20 to 25 :g PM10/m3. When
annual 24 h mean values for PM10/m3 exceeded 30 at Geneva and Lugano, reductions in FVC
and other parameters became apparent. This could be interpreted as possible evidence for a
threshold, since effects with a negative impact for health were observed in the larger cities. On
the other hand, this could also mean that the sensitivity of the test (lung function) was
insufficient to detect effects of inflammatory response or other impacts on respiratory tissues
that have been found to occur in laboratory studies.
In a study that made no attempt to analyse actual exposures to ozone, Galizia and Kinney (1999)
have shown an association between long-term exposure to ozone and reduced lung function in
young adults. They set criteria for new university students they enrolled in their study on the
basis of length of residence ($4 yrs) in a geographical area where ozone levels had exceeded the
summer-season daily 1-hour maximum ozone level of $80 ppb. Among nonsmoking students
from high ozone areas, lung function was generally reduced (FEV1 -3.1% [CI -0.2% to -5.9%]).
FEV25-75 and FEV75 were also reduced. A greater part of this difference was due to the
reductions of lung functions in males. These students also had an overall increase in a number
of respiratory disease criteria including cough (3.1%), phlegm (10.4%) and wheeze (21.5%).
emission.
While respiratory symptoms were similar for males and females, there was higher incidence of
asthma in all males (17.4% compared with females (14.7%). The asthma sufferers were
responsible for a substantial portion of all respiratory symptoms reported by the group. This
study specifically set out to show the separate roles for ozone and particulate matter in chronic
respiratory health effects using the tendency for ozone and PM to co-vary on broad geographic
scales (Galizia and Kinney, 1999). The next step (as yet incomplete) for the analysis was to
include more complete exposure and monitoring data to determine the independence of ozone
induced effects from ambient PM on this group.
5.2.6
Sensitive Populations for Respiratory Impacts
Children represent the population most sensitive to air pollution because of (1) differences in
pulmonary anatomy and physiology (breathing patterns differ between adults and children); (2)
increased susceptibility to respiratory infections; and (3) greater susceptibility to chemical insult
at different stages of development (OMA, 1998).
The epidemiological results point to several sub-populations as being at special risk for PM
exposures: children, individuals (but especially the elderly) with respiratory diseases (e.g., chronic
obstructive pulmonary disease, acute bronchitis, asthma, pneumonia) and cardiovascular disease
(e.g., pneumonia), and the frail (as defined above). As previously noted, a significant portion of
the adults at special risk, such as those with chronic obstructive pulmonary disease, chronic
bronchitis, and cardiovascular disease, suffer from these diseases because of cigarette smoking.
This emphasizes the importance of understanding the interaction of multiple risk factors in
evaluating the risks of low-level environmental exposures to PM and other pollutants.
emission.
6.0
SOURCES OF PM AND ASSOCIATED HEALTH
EFFECTS
6.1
Are all sources of PM10 equally responsible for health
effects?
The contribution of coarse particle PM10 emissions to mortality has been the focus of recent
papers (Pope et al, 1999; Schwartz et al., 1999; Laden et al., 2000; Schwartz et al., 2002). Timeseries studies of patterns of daily mortality and particulate matter as reflected in measurements
of PM10, have observed statistically significant associations with air pollution even after
controlling for seasonality and weather variables. In certain conditions, or in specific areas, the
statistical confidence of the association between air pollutants and mortality is weak and
considerably less certain. Such is the case of the region around Salt Lake City along the west
facing mountain range in Utah identified as the Wasatch Front. Pope used an additional factor
in his analysis that was termed the clearing index. Elevated levels of PM10 can occur not only
on days with stagnant air, but also when high winds are responsible increased from windblown
dust. Days when inversion events are responsible for local accumulations of particulate matter
showed that 70 to 90% of PM was made up of particles less than 2.5 microns. With days when
a high clearing index were screened out, statistical evaluation of some ten years of mortality data
for Salt Lake City revealed the anticipated relationship between PM10 and mortality (lagged by
5 days, using a moving average concentration). The sensitivity of the association between
mortality and PM10 and the reliance on weather patterns is not unusual. The estimated pollution
effect was substantially larger when days with a high probability of wind-blown dust were
excluded, and when a constructed mean value for a number of monitors was used instead of a
single, centrally located monitor (Pope et al., 1999a).
Schwartz et al. (1999), have also prepared a re-analysis of PM associated mortality data using only
days with significantly elevated levels of crustal-derived particles. Like Pope et al. (1999) they
directly suggest that the primary and secondary fine combustion sources of particles commonly
found in stagnant air masses pose the greater health risk. Schwartz et al. (1999) however offer
no indication as to the likely source of these combustion products. This analysis examined days
between 1989 and 1995 in Spokane, WA characterized as episodes of high PM10 concentration
with the dominant feature of coarse particulate from crustal sources. Previous time-series
studies estimated that an incremental increase in PM10 exposure of 221 :g/m3 should be
associated with approximately a 20% increase in daily mortality. During periods of dust storms,
when coarse particles were identified as the predominant source of PM, Schwartz et al.(1999)
found no evidence of increased mortality, even though these qualified as high risk days.
Thus, the re-analysis of the statistical associations for mortality (Pope et al.,1999a; Schwartz et
al., 1999) appear to support the findings of Chow et al. (1999). PM10 data, especially as they
relate to potential health effects and mortality, are (1) most reliable in conditions of low wind
speed and suggest that local conditions for PM10, and (2) are especially sensitive to confounders
such as local activities and interference form fugitive dusts. This has significant implications for
emission.
management approaches to remediation of PM, since it reinforces the connection between
source and the type of emission most likely to be associated with health effects. Because there
are day-to-day correlations between pollution, season and weather variables, confounding of
associations with mortality by weather variables is a concern (Pope et al., 1999a).
In a large study of six U.S. cities, Laden et al. (2000) used the elemental composition of sizefractionated particles to identify several distinct source-related fractions of fine particles. They
then examined the association of these fractions from different sources with daily mortality in
each of the six cities. Combined effect estimates were calculated and they found the following:
• A 10 :g/m3 increase in PM2.5 from mobile sources (i.e., traffic) accounted for a 3.4%
increase in daily mortality
• A 10 :g/m3 increase in PM2.5 from coal combustion sources accounted for a 1.1%
increase in daily mortality
• A 10 :g/m3 increase in PM2.5 from crustal particles was not associated with daily
mortality.
The authors concluded that combustion particles in the fine fraction from mobile and coal
combustion sources are associated with increased morality, but fine crustal particles are not
(Laden et al., 2000). Other studies doing source-oriented evaluations of PM components have
found similar results (Mar et al., 2000; Tsai et al., 2000; Claiborn et al., 2000; Janssen et al., 2002).
These reports suggest that it is primarily combustion-related sources, including secondary
aerosols (sulfates and nitrates), automobile emissions, coal combustion, oil burning and
vegetative (biomass) burning that are associated with excess risk for mortality. In contrast,
crustal particles were generally not positively associated with mortality, with some studies even
reporting negative associations (Mar et al., 2000; Janssen et al., 2002). For example, Janssen et
al. (2002) reported a negative association with PM10 emissions from fugitive dust but strong
positive associations with PM10 from highway vehicle emissions, particularly diesel emissions.
6.2
What is the relationship between traffic and evidence for
adverse health effects from PM?
In Seattle, combustion products that contribute to fine PM from a number of sources including
residential wood burning have been reduced over the past ten years. This has not had any
measurable effect on the incidence of asthma in the non-elderly population. The persistent
association between concentrations of both PM and CO (lagged one day) were related to
automobiles (Sheppard et al., 1999). An interquartile change in PM of 19 :g/m3 (PM10), or 11.8
:g/m3 (PM2.5) lagged one day, resulted in an estimated increase of 4-5% in asthma admissions
to hospital emergency departments (The interquartile range = the population between the 25th
percentile and the 75th percentile. Therefore, the interquartile change of 19 :g/m3 indicates
that when 50% of the observations increase by this concentration, asthma admissions will
increase by 4-5%).
emission.
English et al. (1999) examined the relationship between residence, traffic and health for
asthmatics in San Diego, California. The purpose of the study the effect of motor vehicles on
health by comparing frequency of access of the medical system by a large group of children
diagnosed as asthmatic with similar children with non-respiratory related ailments. There was
no significant association between any health endpoint and traffic. No attempt was made to
model vehicle emissions to obtain a value for exposure. A majority of children (75%) lived
within 550 ft (165 m) of a busy road. There was no evidence of increased asthma risk with
higher traffic counts. Among those diagnosed as asthmatic, and who lived close to the busiest
streets there was a moderate increase in the frequency of access to health care, but this was not
considered significant.
The association between vehicle exhaust and chronic respiratory symptoms in children living
near freeways has been investigated by van Vliet et al.(1997). The respiratory health of girls
rather than boys seemed to be more sensitive to the amount of vehicle traffic within a 1 000 m
radius of their residence. The study assessed lung function as well as chronic respiratory
symptoms for school aged children in South Holland. The study drew associations between the
residence location and the volumes of traffic on adjacent, highly trafficked (80 000-150 000
vehicles/day). Ambient levels of PM10, PM2.5, BS, and NO2 were available from local
community monitoring data. There were significant associations found between girls living
closer to a highway (100 metres) and self reported complaint of chronic cough and wheeze.
There was also an association for chronic respiratory problems for those with diagnosed asthma,
and density of traffic and black smoke detected in the schools. These associations were
generally greater for truck (diesel) traffic than for automobile traffic (van Vliet et al., 1997).
In Tiawan, the association between monitored levels of SO2, NOx, O3, an PM10 and self reported
and physician diagnosed asthma was investigated for approximately 1.02 x 106 middle school
students (Guo et al., 1999). Asthma prevalence varied by district, but was between 4.6% and
12% of students, and the relevant ambient monitoring data for any individual was never more
than 2 km from their school and residence. There were no significant associations between
individual pollutants and asthma for either boys or girls; however, there were extremely strong
associations for three factors including non-summer temperature, humidity in winter and traffic
related air pollution. The latter was defined through principle component analysis that showed
NOx and CO were strongly associated with traffic, but not O3. SO2 and PM10 were more closely
associated with stationary industrial/power generating utilities. The lack of personal exposure
data restricts the ability to interpret this result. While CO correlated with asthma, there is no
mechanistic/clinical data to support this association. The association with wet or damp periods
of the year suggested traffic (road dust/ emissions) and house dust mites were likely the primary
associations with increased episodes of asthma (Guo et al., 1999)
A number of studies have shown an increased association of hospital admissions for children
with respiratory disease, especially asthma, and their places of residence in urban areas
characterized by high traffic volumes. Miguel et al. (1999) surmised that paved road dust
contains allergens deposited on their surfaces, and that these become airborne by virtue of
emission.
traffic volume, increasing atmospheric allergen concentrations to levels above what would be
experienced without motor vehicle traffic. Paved road dust was vacuumed directly from road
surfaces. Collected material was size fractionated to match (artificially) PM10 and TSP samples
of the type comparable to that assessed in respiratory health studies. Comparisons for trace
element and metal analysis made with ambient monitoring data confirm the similarity of
pavement-derived samples with actual local PM10.
Airborne particle (PM10) extracts were found to be more allergenic per unit mass extracted than
paved road dust. Ratios of PM10/Road Dust varied from 100 in Los Angeles to between 4 and
12 at other sites. Calculations presented show that from between 0.5% and 12% of total
airborne allergenicity detected was likely to derive from resuspended road dust (Miguel et al.,
1999). It was noted that significant quantities of airborne particulate matter contains reintrained
road dust in urban areas, and that there are significant levels of allergenic substances in both
road dust and inhalable particulate. Allergens could be fragments of pollens and plant materials
that are stable in the environment, and although they may not be exactly similar to known
allergens, they have sufficient cross reactivity to stimulate allergic responses in susceptible
individuals (Miguel et al., 1999).
emission.
7.0
EXPOSURE ASSESSMENT
7.1
Personal Exposure
Personal exposure is only indirectly related to ambient air pollution, and not all sources exposure
to environmental toxics are attributable to fossil fuel combustion.
Regardless of study design, pollution exposure assessments introduce measurement errors
(Zeger et al., 2000). Few ecologic time-series analyses have properly addressed the question of
quantification of personal exposure. The use of GIS data to precisely locate populations of
interest such as residence in respect of a busy road has been reported in several studies (see.
Brauer et al., 2002). Great reliance is placed on standardized approaches that have not greatly
changed in the past five or ten years (e.g. U.S. EPA PTEAM studies by Özkaynak et al.,1996).
Lipfert and Wyzga (1995) have suggested that the central monitoring data used in the time-series
analyses have uncertain relationships with the exposures of individuals in study communities.
Time-activity analyses suggest that North Americans spend most of their lives indoors (87.2%),
in or near a vehicle (7.2%) but just under two hours (5.6%) outdoors (Valberg and Watson,
1998). This is much more important in the case of Canadians who spend an even greater
proportion of the time indoors. Actual personal exposures and those exposures represented by
central monitors values show considerable discrepancies (Guo et al., 1999; Brauer et al., 1999;
Alm et al., 1999; Boudet et al., 1998; Lioy et al., 1999; Janssen et al., 1998a; 1998b; 1999).
Defining the average personal exposure to PM requires consideration of a combination of
factors. These include true outdoor ambient levels as well as indoor sources and levels. Indoor
sources include tobacco smoke, pets, biogenic sources (dust mites, fungi and bacteria) heating
and cooking and others, as well as those contributions made by the penetration of outdoor
ambient pollution indoors (Zegler et al., 2000). Therefore, the average personal exposure for
a population is related to ambient air pollution, but is offset by the effects of the population
average from non-ambient indoor sources.
Indoor and outdoor ambient particles are not identical in composition (Zegler et al., 2000). The
ratio of fine particulate from outdoors that is generally found indoors is less than one based on
sulphate tracer studies. Therefore, effects based solely on ambient pollution levels would tend
to be revised downwards by whatever fraction of adverse health effects can be attributed to
particulate and other matter (e.g. VOCs) that can be ascribed to indoor sources.
A number of sources of uncertainty can be identified for any study that attempts to attribute risk
for health effects to a population. The expression of uncertainty takes many forms, but the one
we are most aware of are the 95% confidence intervals that accompany any statistical
association. These merely mean an effect can be assumed to fall within some expressed range
of values at least 95% of the time.
emission.
Uncertainties introduced through measurement error is the easiest source of uncertainty to
correct (Zegler et al., 2000), but measurement errors can not be said to contribute sufficient
uncertainty as to make time-series models uninterpretable. The largest biases introduced for
inferences for health effects will be for the mortality-personal exposure relative index because
of errors between ambient and average personal exposure (Zegler et al., 2000).
When data from PTEAM studies (Özkaynak et al., 1995) for personal exposure indoors have
been used, the coefficient of association with health effects are greater that if the regressions
used rely on estimates from ambient pollutant levels (Zegler et al., 2000). This would suggest that
personal indoor exposure have a greater potential impact for health than do outdoor sources. Therefore, differences
between average personal exposure and ambient measurements are the most likely source of substantial bias in
time-series studies. Wilson and Suh, (1997) have conducted a meta-analysis of data from multiple
sites and concluded that the concentration of particles that originate from indoor sources are
independent of ambient levels over time. This may be particularly important for estimation of
mortality risk because it would give greater weight to particulate exposures from indoor sources.
In summary, improved monitoring programs with better information about the variation in local
communities across an urban area will tend to improve predictions for health effects, but would
not be likely to change the trend in health effects observed in time-series studies. Sources of
bias from ambient measurement uncertainty is smaller than personal exposure bias.
7.2
What are the relationships between outdoor ambient
monitors or predicted levels of PM or VOC and actual
human exposure?
Adequate exposure models are essential to evaluate the association between exposure
concentration and dose deposited in the lung. One of the major factors that determines the
toxicity of inhaled particles is their dosimetry characteristics and deposition rates (Salvi and
Holgate, 1999). Also critical for adequate risk assessment are associated rates for clearance,
retention, translocation and dissolution which may be very different in different regions of the
lung. Deposition efficiencies of inhaled particles in the human respiratory tract has been
modelled using carefully designed replicas of an adult and a child’s upper respiratory tract
(Oldham et al., 1997). They confirmed previous findings that greater deposition rates in upper
air ways occur in children. Although results matched predictions, and significantly greater
amounts of particles deposited in the trachea and first and second bronchial arches of a 4-7 year
aged model than when the same sized particles were administered to an “adult”, particle
behaviour and deposition rates could not be adequately predicted beyond these areas (Oldham
et al., 1997).
Personal exposure to PM was evaluated in different activity-related exposure scenarios (Brauer
et al., 1999). Samples were collected using a personal monitor with a PM2.5 impactor. Highest
exposures occurred during commuting in any diesel powered form of transportation (bus, train,
ship), but was unaccountably low for bicycle couriers in Vancouver. A variety of activities were
emission.
performed including commuting by different modes of travel, outdoor and indoor activities.
Some significant specific sources of exposure were indoor cooking, environmental tobacco
smoke and wood burning. In concert with earlier observations that employed personal
monitoring methodologies, Brauer et al. (1999) noted a “personal cloud” which was attributed
to local resuspension of dust. This personal cloud increased local PM concentration by more
than 50% compared with samples collected coincidentally outdoors, or at another site indoors.
Carbon monoxide and PM have been monitored inside an automobile in a medium sized town
(pop. 85 000) in Finland (Alm et al., 1999). Automobile commuters are exposed to higher
concentrations of PM than bus and rail commuters. VOC and CO exposure for car drivers was
higher than for a bicyclist. CO concentrations and exposures were generally greater in the
morning than the afternoon. Variable PM counts were reported, the highest being associated
with heavier traffic and colder temperatures. There was no seasonal effect for driver exposure,
and generally slower speeds increased particle exposure rates.
7.3
Personal Exposure Monitoring and PM
Current monitoring for particulate matter undertaken by the MOE in Ontario has been generally
changed to measure PM2.5. This has not always been the case, and much of the health evaluation
has depended on interpretations using PM10. In Canada, FP averaged across all sites comprises
49% of PM10 (Brook et al., 1997). In Toronto the ratio of FP to PM10 ranges from 0.15 to 0.90
for the 5th and 95th percentile respectively. The interquartile range (the difference between the
25th and 75th percentile) at Toronto is close to 0.25. For most sites in Canada, however, the ratio
of PM2.5 to PM10 at a given site varies only ± 0.10, fifty percent of the time (Brook et al., 1997).
Boudet et al.(1998) attempted to carry out a source apportionment study for daily personal PM2.5
exposure using personal samplers and detailed activity report logs. Thirty-three percent of total
personal exposure based on mass could be accounted for aggregate time spent outdoors,
although only 11% of time was spent outside (Boudet et al., 1998). The average indoor
(personal) exposure was 7.5 :g/m3, and the outdoor exposure was 29 :g/m3. They found that
while co-located samplers operate with similar efficiency, there were large variability problems
and fluctuations in measured personal exposure. Commuting time may represent a larger
portion of total PM exposure than initially realized by a simple correction for “time spent
outdoors”.
Time-activity analyses suggest that North Americans spend most of their lives indoors (87.2%),
in or near a vehicle (7.2%) but just under two hours (5.6%) outdoors (Valberg and Watson,
1998). Elderly individuals with preexisting respiratory or cardiovascular disease have been
repeatedly determined to be at greater risk from ambient PM. Among this segment of the
population, the greater balance of life is spent in buildings or interior spaces, and not in direct
contact with the outdoors. Thus, the indoor environment is the major exposure category, and
daily fluctuations of indoor air quality are the primary concern for association of ambient
conditions with specific health outcomes (Valberg and Watson, 1998). A linkage must be made
emission.
between remotely monitored PM, evidence of adverse health impacts based upon biological
plausibility and epidemiological analysis, and the interior spaces where most of the population
is to be found. In a recent study of the indoor/outdoor relationships for ambient PM2.5, sulphur
dioxide was highly correlated indoors and outdoors, but indoor concentrations average only 3%
of ambient SO2. Indoor penetration of HNO3 is substantially less than sulphate penetration.
Fine particulate mass had a low indoors/outdoors correlation in this study, and was more closely
related to the period when indoor spaces were occupied by people (as a source of fine PM)
(Patterson and Eatough, 2000).
Fine particulate matter (FP) in the outdoor environment in particular appears to contribute to
indoor air pollution, but recent experience of personal monitoring studies has recorded higher
personal concentrations of FP even though the indoor levels may be below outdoor ambient
concentrations (Janssen et al., 1999). Since the lung does not discriminate among fine particles
on the basis of source, it is important to characterize sources for possible health effects. Lioy
et al. (1999) have shown that micro-environments in buildings can be responsible for large
fluctuations in indoor particulate matter. These variations may be due to, for example, vacuum
cleaners, and whether exhaust air from the appliance receives HEPA filtration. The brushes of
vacuum cleaner motors were found to produce aggregated particles of carbon (<0.3 :m) during
operation. Vacuum cleaner bags also performed poorly, and were found potentially responsible
for the redistribution of biogenic/allergenic materials. Some of these have been implicated as
causative factors for asthma (Habbick et al., 1999).
It is important to demonstrate this relationship if we are to accept the use of outdoor
concentration as a surrogate for personal exposures. Living near a busy road, and time spent
in a vehicle have been found to be the greatest sources of variation in personal exposure levels.
Janssen et al. (1998a; 1998b; 1999) have addressed questions of personal exposure, and the
relationship between indoor and outdoor fixed monitors for PM10 and PM2.5. Janssen et al.
(1998a) were able to validate fixed site monitors as measures of personal exposure by
comparison of concentrations of fine particulate matter measured with personal samplers
(4L/min, >3:m cutoff) and measurements at stationary monitors at the same location. Janssen
et al. (1998b) found that for an individual adult, the correlation between exposure assessment
from a personal monitor and a fixed monitor was quite high. Outdoor concentrations of PM10
at fixed monitors usually exceeded PM10 found at fixed monitors indoors, but outdoor air
considerably underestimated the personal exposures of adults monitored in this group. Even
though none of the subjects smoked, even occasional or incidental exposure to environmental
tobacco smoke (ETS) interfered with the correlation of indoor to ambient outdoor
measurements. On average, personal air monitors registered 20 :g/m3 over outdoor ambient
levels, and indoor stationary monitors registered 7 :g/m3 under outdoor ambient PM10. The
existence of a “personal cloud” that contributes to personal PM measurements was reported by
the PTEAM study (Özkaynak et al., 1996).
In another study, fine particulate monitoring for exposure to school-age children showed
remarkable correlations between indoor and outdoor concentrations of fine mode particulate
emission.
(FP distribution is generally between 1:m and 3:m) (Janssen et al., 1999). ETS added
significantly to the personal monitoring results of children. This was reflected in a poor
correlation of 0.41 for FPpersonal/FPambient. The crossectional correlations for FP among the
children in the study were very high (R = 0.82), probably indicating that children who have a
shared classroom experience, live in similar neighbourhoods and even in some instances in
houses with similar structural configurations show a reduced inter-individual variation. This
study was probably not adequate for purposes of demonstrating differences in exposure across
a large population, but for children engaged in similar organized activities such as school, and
living in similar neighbourhoods, personal exposures would be similar. As with PM10 results in
adults, FP exposure by personal monitoring was greater than expected from ambient
monitoring.
FP does not settle out, so activity does not alter the exposure concentration for fine particles
(Özkaynak et al.,1996; Janssen et al., 1999). PM2.5 is subject to long range transport, and is evenly
distributed over a community (Burton et al., 1996; Wilson and Suh, 1997), therefore dependence
on a single monitor for ambient fine particulate matter (PM2.5) is adequate for personal exposure
assessment (Janssen et al., 1999).
Özkaynak et al.(1996) found that the composition of FP indoors is similar to outdoors, and
suggests this is a reason that the epidemiological studies are able to correlate daily mortality with
outdoor PM. If the most biologically active portion of the PM outdoors is due to the FP, then
outdoor may function as a surrogate for exposure. The fundamental finding that is continually
raised to support this assumption is the relatively invariant content of sulphate found in samples
of indoor and outdoor FP. Yet it is clear that additional (especially allergic or biogenic)
components primarily associated with indoor sources can be shown to be responsible for
serious illness (Habbick et al., 1999).
In summary, we can accept the values for PM observed at fixed ambient monitors as a
reasonable estimate of personal exposure, but we should remember that there is a large amount
of variation between what is registered at the personal level and what is apparent in outdoor
ambient conditions. A significant degree of fluctuation can be introduced by local activities, and
these are not always shared by the entire community.
8.0
INTERACTION AMONG POLLUTANTS
MEASUREMENT OF EFFECT ON HEALTH
8.1
Interactions of Pollutants and Allergens
AND
The presents of allergens and biogenic material in samples of PM is well recognized. Unlike
PM10, PM2.5, SO2, NO2, metals, etc. there has been a weak effort at characterizing the biological
component or matter in particulate. Thus the basis for analysis of interactions between PM and
allergens is not well established (Anderson et al., 1998). Efforts to demonstrate associations
between ambient levels of urban air pollutants and alterations of the immune functions have
been only moderately successful. The problem for assessing individual exposure to particulate
emission.
matter and the linkage of exposure to individual health assessment has not yet been overcome.
The use of central monitors to evaluate exposure is inadequate for exposure assessment
(Koenig, 1999).
These reports suggest, but are somewhat inconclusive in identifying a role for mobile emissions
in reinitiating a preexisting disease condition of asthma. It is clear that meteorological
conditions play a large role in creating the conditions that can produce the potential for health
effects such a asthma exacerbation.
8.2
Air Pollution and Asthma
Mechanical methods, primarily spirometry, have been widely used to measure changes in lung
function in epidemiology studies designed to determine the effect of pollution on populations,
and to characterize disease-related effects of asthma and other respiratory diseases (Table 5).
Controlled exposure studies have shown that inhalation of SO2 (1.0 ppm) for 10 minutes during
moderate exercise can reduce FEV1 by 23%, and increase total lung resistance by an average of
67% (Koenig, 1999). Moderate exposures (0.2 ppm, 6 hrs) to ozone, on the other hand does
not generally produce significantly different effects on lung function in asthmatics, but O3 is a
risk factor in asthma sufferers for increases in bronchial hyperresponsiveness (BHR). Shorter
exposure times to higher levels of ozone (0.4 ppm, 2 hrs) does have a greater affect on lung
function measurements in asthmatic patients (Basha et al., 1994).
Controlled exposures to particulate matter (PM) have been difficult to achieve. More recent
efforts have focussed on the use of concentrated particles recovered from ambient air to
produce controlled, experimental exposures to PM. Results of such exposures have generally
failed to demonstrate significant effects in normal subjects, and there have been no reported
responses of asthmatics to controlled PM exposures (Sioutas et al., 1995).
Epidemiology studies that attempt to demonstrate changes in lung function in response to
pollution can suffer from a number of reliability problems. Generally, these have relied on self
reported changes in peak expiratory flow (PEF) (Table 5). The necessity to accept data collected
in unsupervised conditions is a recognized confounder for such studies. Nevertheless, a number
of studies have recorded differences in lung function parameters between asthmatic and healthy
subjects in response to ambient levels of air pollution. Recently, Mortimer et al. (2001)
concluded that peak flow monitoring performed by asthmatic children used as a sole
measurement tool was not a strong predictor of utilization of medical services within a period
of thirty days. Peak flow values reflect clinical status at the moment the measure is taken, while
symptoms reflect the perceived condition persisting throughout the day. On the other hand,
parents, noting symptoms could encourage a child to use medicine to obtain relief of symptoms.
Thus PEF adds little information that is not already available through an indication of
symptoms in response to a questionnaire.
In Seattle, a change of 11 :g/m3 of PM2.5 was associated with a 15% increase in ER visits, while
emission.
a similar increase (10 :g/m3) resulted in a 3 to 6% increase in ER visits in Anchorage, AK
(Koenig, 1999). Some studies have found a correlation between CO and asthma-related ER
visits, while others have not (Gordian et al., 1996; Yang et al., 1997). A London (UK) study
found ER visits were associates with SO2, NO2 and PM10 (Atkinson et al., 1999). Attribution
of effects to biogenic sources of material, and assigning responsibility for asthmatic respiratory
responses has only recently begun to be considered (Anderson et al., 1998). Statistical
correlations between pollutants and respiratory disease endpoints have been described for
ozone, oxides of nitrogen and sulphur as well as PM, but no consistent pattern is evident.
Different age groups appear to respond differently to different pollutants with varying lag times.
There does not appear to be a single pollutant, or unique group of pollutants or endpoints that
successfully characterize the relationship between increased numbers of ER visits and urban air
pollution (Koenig, 1999).
8.3
Hospital Admission Studies for Asthmatic Patients
Admissions to hospitals for treatment of asthmatic symptoms has been studied in Western
Europe (Barcelona, Helsinki, London and Paris) (Sunyer et al., 1997), where an increase of 50
:g/m3 in NOx is linked to an increase in hospital admissions (1.029 [CI 1.026- 1.055]) for both
sexes age 15 to 64. Under age 15, 24hour levels of SO2 and NO2 lagged three days are
associated with increased hospital admissions for asthmatic patients (SO2 1.075 [CI 1.026-1.126]
and NO2 1.026 [CI 1.006-1.049]).
Sheppard et al. (1999) examined non-elderly asthma patients in Seattle, Washington for
associations between hospital visits and urban air pollutants. They found a significant
association between elevated PM10 lagged one day for an interquartile range (IQR) of 19 :g/m3
OR=1.05 [CI:1.02-1.08]. Overall, there was a 4-5% increase in the rate of hospital admissions
with a change in PM, and a 6% increase with a change in CO or Ozone. A similar association
was found for PM2.5 at 11.8 :g/m3 OR=1.04 [CI:1.03-1.07]. Carbon monoxide (924 ppb) and
O3 were also associated with increased hospital admissions of asthmatics in the population (CO
OR=1.06 [CI 1.03-1.09]; and O3 OR=1.06 [CI 1.02-1.11]). There was no evidence of an
association with SO2 The persistent effect of automobiles on health may be inferred from the
association with CO, since there is no mechanism that links asthmatic respiratory response to
CO.
In another Seattle study, Norris et al., (1999) was able to demonstrate an association between
fine particles and asthma emergency department visits for children. Daily emergency
department counts were regressed against PM, CO, SO2 and NO2. A change of 11 :g/m3 in
fine PM was associated with a rate of 1.15 [CI 1.08-1.23]. PM and light scattering measurements
were highly correlated. CO was also correlated with PM, but not NO2 or SO2. The average
number of visits to an inner city emergency department for asthma by children under 18 years
was 1.8 per day, with a maximum of 9 visits on any one day. The greatest concentration of PM
in urban air was observed in the winter months. Because CO has no biologically plausible
mechanism for exacerbation of asthma, its association was interpreted as a general indicator of
emission.
the stagnant conditions associated with high pollution days in urban areas. PM2.5 in Seattle was
influenced by (a) incomplete combustion consisting of CO, elemental carbon, organic carbon
and soluble potassium (a wood smoke marker); (b) secondary aerosols consisting of ammonium
and sulphate; and (c) fine and coarse crustal material.
In summary, there is growing toxicological evidence for a generalized adverse effect of air
pollution and specifically PM10 the respiratory health of children (Table 5). There is some
uncertainty surrounding these results because of the use of somewhat subjective clinical
measures and questionnaires to assess health impacts. Evidence collected from examination of
specific endpoints recognized as indicative of reduced lung function is most persuasive where
the greatest difference is apparent for opportunities for chronic exposure to elevated PM10.
Jedrychowski et al. (1999) was able to show a significant difference in spirometric parameters
only among young males in highly polluted urban locations. Boezen et al. (1999) and Heinrich
et al. (1999) were also unable to characterize adverse effects on lung function of children living
in conditions of greater exposure to PM10. In agreement with earlier findings, Boezen et al.
(1999) did find evidence of adverse health outcome for children with pre-existing respiratory
disease (allergy and asthma).
emission.
9.0
PLAUSIBLE MECHANISMS
9.1
Biological Plausibility: Linkage between particulate matter
(PM) and adverse biological effect
Guiding principles that address the relationship of markers of deteriorating air quality with
health related endpoints have been developed (HEI, 1999). These principles apply to virtually
all research that poses questions of toxicological linkage between human respiratory and cardiac
diseases and substances found in urban air. The interpretation of experimental animal studies
for purposes of accurate extrapolation across species to humans is dependent on many factors
including:
• The exposure level (the concentration of each pollutant a person or animal is exposed
to in different microenvironments and the time spent in these microenvironments);
• The internal dose (amount of a pollutant or its metabolites in the body after exposure);
• The biologically effective dose (the amount of pollutant or key metabolites that reaches
important target sites in the body);
• The molecular effects (e.g., changes to the structure or function of molecules);
• Other biological responses potentially associated with the development of clinical
disease, such as cancer (formation of some specific DNA adducts, mutations, and
chromosomal aberrations).
Key tools for effect extrapolation are biomarkers, defined as measurable, quantifiable biological
indicators detectable in breath, biological fluids, or tissues. Biomarkers provide a useful means
to evaluate exposure, absorbed dose or dose delivered to critical target sites, to identify
metabolic pathways, and to associate molecular events with specific chemicals or their
metabolites. Biomarkers can also be defined in terms of selected health endpoints, or as
indicators of susceptibility to disease (HEI, 1999).
Plausible mechanisms for PM related morbidity, and especially for exposure to low
concentrations of PM2.5, are the subject of much debate. Effective strategies for minimizing
health risks of PM exposure require the development of mechanistic linkages that extend from
sources of various PM constituents to the development of human health effects. Scientific
evidence from animal studies suggests several possible mechanisms that could establish a
relationship between exposure to quite low concentrations of PM2.5 and the increase in the
several indices of respiratory and cardiovascular health found to be associated with PM
exposures in the epidemiological studies. The absence of any strong mechanistic underpinning
for how PM2.5 may be causing adverse effects results in defaulting to a view that all particles in
the PM2.5 fraction irrespective of their size and chemistry are of equal toxicity per microgram
of mass.
There has also been speculation about the role of various specific chemical species in triggering
these responses and, ultimately, in increased morbidity and mortality. These chemical species
emission.
include acidic compounds, trace elements, and reactive species.
9.2
Inflammatory Response
Much of the data in support of biological plausibility for a linkage between PM and health is to
be found in the recent research literature from investigators who have set out to characterize
the relationship between exposure to particulate matter and toxicological response. The role
of particulate matter in the initiation of the inflammatory response has been examined by a
number of investigators. There is clear evidence that a number of steps are involved in the
initiation of the inflammatory response, suggesting the necessity to overcome a threshold.
Veronesi et al. (1999) used a human bronchial epithelium in vitro to show how residual oil fly ash
(ROFA) particulate could initiate an inflammatory response. ROFA is emitted from residual
oil burning electrical generating facilities as acidic FP (<0.2 :m) that is >95% soluble in water.
ROFA exposure in rodents by intsillation produces hyper-responsiveness in airways and acute
lung injury including epithelial damage, pulmonary oedema, haemorrhage and an influx of
neutrophils (Dye et al., 1997; Kodavanti et al., 1998). In rodent macrophage preparations, ROFA
exposure shows evidence of oxidative damage (reactive oxygen species), release of cytokines,
and initiation of apoptosis. Similarly, human macrophages treated with ROFA particulate show
bursts of oxidative metabolism and apoptosis. Many of the inflammatory responses have been
associated with the presence of soluble transition metals such as nickel, vanadium (4+, 5+) and
iron (3+) and sulphates in the acidic particulate (Kodavanti et al., 1998). Veronesi et al. (1999)
showed the combination of strong acidity with transition metals present could initiate cytokine
expression, and was completely dependent on the acid mediated destruction of bronchial
epithelium. Co-exposure to an antioxidant led to the abolition of the transition metal related
effect.
Adamson et al. (1999a, 1999b) have investigated the pulmonary toxicity of an urban atmospheric
particulate sample (EHC-93) collected in Ottawa, Canada. It was concluded that the toxicity
in the whole urban air particulate matter (PM4.6) was due to the water soluble fraction. This
contained sulphate and transition metals, but was not particularly acidic. Adamson et al.(1999a)
conclude that their study supports the soluble fraction of urban dust as having the potential to
cause lung cell injury as well as inflammation.
Cytotoxicity and induction of proinflammatory cytokines from human monocytes has been
reported by Monn and Becker (1999). Cytokine induction (IL-6) in monocyte cultures was
primarily activated by the water soluble fraction from coarse particulate (PM10-2.5) that was
collected outdoors. This IL-6 cytokine responses are related to biogenic sources such as
endotoxins. Enzymatic destruction of the biogenic material reduced the endotoxin mediated
cytokine response. Monn and Becker (1999) concluded that cytotoxicity and induction of
cytokines are associated with the coarse fraction from outdoors, and that coarse or fine
particulate from indoors was inactive. Fine particles from outdoors did not activate cytokine
production.
emission.
In another study using the urban air particulate matter EHC-93, Adamson et al. (1999b) have
provided additional support for a toxicological response to PM. These studies involved the
administration of large doses of inhalable PM4.6 (5 to 57 mg/m3) in conjunction with ozone (800
ppb) to rats. Inflammation and recruitment of polymorphonuclear leukocytes occurred at the
site of injury whether produced by dust (high dose) or ozone alone. Lower doses of particulate
did not elicit a measurable response in vitro suggesting a possible threshold for this response.
The greatest inflammatory response was observed by Adamson et al. (1999b) when both agents
were present.
Nasal epithelium is sensitized to environmental insults from urban air. Calderón-Garcidueñas
et al. (1998) examined children residing in an area suffering frequent episodes of high pollution.
Damage in the nasal epithelium shown by biopsy show evidence of pollution related to nasal
squamous metaplasia and dyspnea. Cellular changes observed in nasal tissues of residents of a
community in south west Mexico City were correlated with prolonged exposure to elevated
levels of ozone, formaldehyde, acetaldehyde and PM.
emission.
10.0
TOXICOLOGICAL EFFECTS OF DIESEL EMISSIONS
10.1
Introduction
An excellent review of the relationship between diesel emissions and asthma mechanism of
inflammatory response is available in the literature (Pandya et al., 2002). Amongst the various
sources of fine particulate matter, diesel exhaust is the single major contributor in most urban
areas. Urban particulate matter is composed of emissions from many sources. In the United
States, the California Air Resources Board has declared diesel particulate emissions to be toxic
(CARB, 1998). They have identified both a cancer and non-cancer risk specific dose. The
United States Environmental Protection Agency has recently released a health assessment
document for diesel engine exhaust (U.S. EPA, 2002b). This document concluded that there
are acute and chronic non-cancer exposure effects of diesel emissions, and it also concluded that
there was evidence for long-term carcinogenic effects of exposure.
When inhaled, approximately 50% of fine PM in diesel emissions would be expected to be
deposited in the lung (Salvi and Holgate, 1999). The mass median diameter of diesel exhaust
is approximately 0.2 :m, with over 90% of particles being less than 1 :m (HEI, 1995). Under
the electron microscope, >80% of diesel PM was <0.1 :m to which a large number of
compounds (~18 000) are adsorbed. In inhalation studies of diesel emissions carried out in rats,
there was evidence of impaired macrophage clearance attributed to the increased burden from
excessive accumulation in the lung tissues after prolonged exposure to fine particulate (Salvi and
Holgate, 1999).
10.2
Non-Cancer Toxicological Effects of Diesel Emissions
Diesel exhaust particles have also been recently implicated in the increased prevalence of asthma
in urban populations. Bronchial asthma may be divided into allergic (extrinsic) and nonallergic
(intrinsic) forms of the disease. Epidemiologic investigations and experimental studies on the
nose suggest that diesel exhaust particles may be implicated in the rising prevalence and
increasing severity of allergic diseases. Allergic rhinitis is notable in school-aged children in
urban areas. Joint administration of allergen and diesel exhaust particles has suggested to
possibility that combustion particles exacerbate or have an adjuvant effect on natural allergic
reactions (Takano et al., 1997).
Oxidative damage mediated by particles deposited in lung tissue leads to activation of
proinflammatory cytokines IL-1, IL-6 and TNF". Instillation of fine carbon particles similar
to diesel emissions, but without the associated polycyclic aromatic compounds (PAC) of
combustion stimulates the release of neutrophils from bone marrow in rabbits (Terashima et al.,
1997). Instillation of diesel particles in mice has demonstrated aggravation of allergen-induced
air way inflammation, increased IL-5 in brochoalveolar lavage fluid, a 35 fold increase in
eosinophilic infiltration and a six fold increase in lymphocytes in lung tissue (Takano et al., 1997).
Fossil fuel combustion products have an adjuvant effect on the immune system. Inhalation of
emission.
diesel derived particulate matter may enhance allergic inflammation reactions. Diesel particles
or organic extracts of these particles affect IgE production. When mixed with known allergens
such as Japanese cedar pollen, diesel particulate can induce IgE production (Nel et al., 1998).
The lack of adequate exposure-response information in the acute health effect studies precludes
the development of recommendations about levels of exposure that would be presumed safe
for these effects (U.S. EPA, 2002b). EPA has concluded that sufficient evidence is available
from human and animal studies to demonstrate that acute or short-term (e.g., episodic) exposure
to diesel exhaust (DE) can cause acute irritation (e.g., eye, throat, bronchial), neurophysiological
symptoms (e.g., lightheadedness, nausea), and respiratory symptoms (cough, phlegm). There
also is evidence for an immunologic effect-the exacerbation of allergenic responses to known
allergens and asthma-like symptoms.
Tables 6 and 7 summarize the most important observations for effects of diesel particulate
matter in laboratory animals and in controlled studies in humans. Controlled exposures of
humans to diesel exhaust (100 :g/m3) in inhalation chambers produced increased numbers of
neutrophils in vivo, and in alveolar macrophage cells isolated from bronchoalveolar lavage, cell
function was reduced when tested in vitro. Air way resistance during exercise was increased after
exposure of human volunteers to diluted diesel exhaust for one hour (Rudell et al., 1996). PAH
or polycyclic aromatic compounds (PAC) associated with diesel exhaust particles can affect
cytokine production in macrophages. This may occur either by direct reaction with PAH or as
a consequence of interaction of their oxidative metabolites such as quinones (Nel et al., 1998).
Metabolic conversion of PAH by CYP1A1 and other cytochrome p450 can increase the level
of reactive oxygen species, known to affect inflammatory responses. Chemokines such as MIP1", MIP-1$; MCP-1, 2 and 3; and RANTES (Regulated on Activation, Normal T-cell
Expressed and Secreted) are produced by macrophages in response to diesel particulate (Nel et
al., 1998).
emission.
Table 6: Adjuvant Effect of Diesel Exhaust Particulate on in vivo
Allergic Responses in Animals
Effect or Response ObservedA
Increased IgE production
Allergen-specific IgE enhancement after intraperitoneal, intranasal, intratracheal, inhalation, or foot pad
challenge when particles are mixed with a known allergen.
Intraperitoneal pyrene extracts of diesel exhaust particulate act as adjuvant for IgE production.
Increased IL-4 production in mediastinal lymph nodes after intratracheal or inhalation challenge.
Increased local inflammatory response in poplietal lymph nodes after foot pad challenge.
Induction of a TH2-like cytokine profile in airways
Diesel exhaust particulate enhance sneezing, frequency and volume of secretions, and airway resistance in a
histamine induced rhinitis model.
Diesel exhaust particulate increase vascular permeability in parallel with increased nasal airway resistance in
guinea pigs.
Enhancement of airway constriction, eosinophilic inflammation, goblet cell hyperplasia after intratracheal or
inhalation challenge.
A
From Nel et al., 1998
Table 7: Adjuvant Effects of Diesel Exhaust Particulate on in vivo
Allergic Responses in Humans
Effects of aerosolized diesel exhaust particulate on nasal allergic responseA
Increased local production of IgE and IgE-secreting cells.
Qualitatively different isoforms of IgE produced through alternative forms of g mRNA
Augmented production of allergen-induced antigen-specific IgE.
Nonspecific stimulation of a broad cytokine profile in the absence of allergen.
Induction of a TH2-like cytokine profile when intranasally administered together with an allergen in the nose.
Interact with allergen to drive isotype switching of B cells to IgE.
Induce an influx of T cells, monocytes and granulocytes, but NOT eosinophils.
Augmented production of C-C chemokines (RANTES, MCP-3, and MIP-1", but not eotaxin) on intranasal
challenge.
Effects of inspired diesel exhaust particulate on the lower respiratory tract
Increased neutrophil number and decreased phagocytosis by alveolar macrophages on challenge with diesel
exhaust fumes.
Increased airway resistance on challenge with diesel exhaust fumes.
A
from Nel et al., 1998).
emission.
emission.
Thus diesel particulate can act directly to initiate inflammatory responses but also on conjunction
with other molecules and allergens on mucosal tissues to enhance inflammatory responses.
Salvi et al. (1999) have carried out human chamber studies to investigate the acute inflammatory
responses in airways to short term exposure to diesel exhaust. Diesel particulate in the
accumulation mode (0.02 to 0.2 :m) have a high surface to mass ratio and therefore are capable
of carrying larger per-particle burdens of metals and organic carbon. Salvi et al. (1999) asked
whether exposure to diesel exhaust might induce inflammatory and mediator responses in
airways. This is a direct observation in humans of the effects recorded in rats using urban
particulate (Adamson et al., 1999b) or ROFA (Kodavanti et al., 1998; and others). Exposures
were for one hour at 300 :g/m3 of PM, 6.1 ppm NOx, 7.5 ppm CO and 4.3 ppm hydrocarbons.
No changes were observed in respiratory function (FEV1, FVC, PERF) of healthy volunteers.
Six hours after exposure, markers of inflammatory response were examined in lung lavages of
the volunteers. Marked increases were noted in neutrophils, mast cells and lymphocytes. It was
concluded that diesel exhaust was positive for expression of events that lead to inflammation, but
these were not greatly elevated compared with an greater acute exposure. The NO2 exposure (1.6
ppm) is reflective of high levels of pollution, but is known not to induce inflammation in this
assay. The authors suggest that the particulate component was responsible for the increase in
markers of inflammation, since these contain metals (Fe3+) known to be capable of the
production of hydroxyl radicals and oxidative stress. Salvi et al. (1999) finally concluded that
standard lung function tests underestimate health effects of particulate exposure, and therefore
epidemiology studies likely underestimate tissue insults from diesel particulate and exhaust.
The measured exposure rate of 300 :g/m3 used by Salvi et al. (1999) was for particles with a
MMAD of 0.02 to 0.2 :m. This is in itself a large exposure, but should be considered in the
context of the urban environment, where other sources of PM would add substantially to this
value. At best, it is an order of magnitude in excess of the recommended AQO guideline for
PM2.5, and would represent a day with exceedingly poor air quality for anywhere in Canada. The
authors state that the levels are comparable to those found “in mines, garages, bus stations
railway stations and congested streets of the developing world”.
Nordenhäll et al. (2001) have shown that when asthmatics are exposed to high concentrations
of diesel exhaust, typical hyperresponsiveness of the respiratory system is detectable in small
airways. Diesel exhaust is a major source of particulate matter in large urban centres such as
London, England where as much as 87% of PM10 particulate emissions has been attributed to
vehicles. A Swedish team recently investigated respiratory responses among asthmatics,
especially after short-term exposures to diesel exhaust. Fourteen adults who exhibited responses
to airborne allergens and who were also diagnosed asthmatics received controlled exposures to
diesel exhaust. Their respiratory responses both immediate and after 24-hours were tested for
(1) hyperresponsiveness of airways, (2) lung function, and (3) airway inflammation. Standardized
diesel exhaust exposures were administered over a period of one hour in an environmental
chamber at a particle concentration of 300 :g/m3 and a median NO2 concentration of 1200 ppb
(1.2 ppm or 2,292 :g/m3 NO2) (Nordenhäll et al., 2001). In terms of comparable exposure
emission.
levels in Hamilton, these controlled exposure conditions were extreme, and could only occur
under highly specific, localized conditions.
In this study of asthmatic individuals exposed to high levels of diesel exhaust particulate and
associated pollutant gases there was no change in lung function parameters FVC and FEV1
immediately following exposure, but there was a small increase in airway resistance. When tested
24-hrs after exposure by a methacholine challenge to evaluate changes in forced expiratory flow,
virtually every subject exhibited reduced lung function compared to pre-diesel exposure.
Therefore, although impacts on lung function were not immediately observed, a delayed
sensitivity to asthmatic/respiratory challenge was detected. Epidemiological data suggests that
health outcomes of exposure to PM are generally recognized after a one to four day lag period,
indicating that the delayed sensitivity observed was to be expected. During the post diesel
exposure period, inflammatory responses including cytokine IL-6 production was initiated
(Nordenhäll et al., 2001).
Researchers have long been aware that rodent exposures to diesel exhaust produce inflammatory
responses in the lung (Watson and Green, 1995). Airway hyperresponsiveness is a prominent
characteristic of asthma. Diesel exhaust particulate (DEP) has been shown to induce airway
hyperresponsiveness in a mouse model by several authors. Takano et al. (1997) were the first to
show responses in ICR mice presensitized to ovalbumin and diesel exhaust particles after
repeated tracheal instillation over a period of six weeks. After the final treatment histological
assessment showed macrophages were more prolific in ovalbumin and ovalbumin + diesel
particulate (OVA+DEP) treated mice, but not in mice that received diesel particulate alone. The
response in the OVA+DEP mice, there was a twenty fold increase in eosinophils in
broncholavage fluid, ten fold increase in neutrophils, and a four to five fold increase in the
number of lymphocytes. Epithelial cells in the air ways of these mice showed evidence of
hypersecretion and an increased number of goblet cells. Lung tissue showed elevated levels of
the cytokines GM-CSF and IL-2. These studies confirm the mouse model (a) mimics the
bronchial asthma which is an inflammatory disorder of air ways, (b) is accompanied by airway
hyperresponsiveness, (c) and clearly show the adjuvant effect of diesel exhaust particles (Takano
et al., 1997; Takano et al., 1998).
These results were confirmed by Miyabara et al. (1998a, b) who used a similar model. Daily
inhalation of diesel exhaust is impractical for mice, but a similar effect can be obtained from
tracheal instillation experiments. In these experiments, mice were anaesthetized, a tracheotomy
performed, and ventilated with a rodent ventilator. Miyabara et al. (1998a) also noted increased
eosinophil and neutrophil infiltration of lung tissue, and the enhanced production of allergen
specific OVA-IgG1. In another, similar experiment, Ohta et al. (1999) demonstrated that it was
unnecessary to sensitize mice with an allergen such as ovalbumin to observe a response to diesel
particulate. They were able to show evidence for increased levels of the proinflammatory
cytokine GM-CSF in response to diesel exhaust particulate in A/J and C57B1/6 mice. GM-CSF
promotes differentiation and activation of eosinophils. In their mouse model, exposure to diesel
particles was achieved by intranasal administration over a period of two weeks. There was no
emission.
evidence of increased IgE production, indicating that air way hyperresponsiveness was not
mediated by IgE. However, mice treated with an antibody that neutralized the proinflammatory
cytokine (anti-GM-CSF) showed virtually no evidence of air way hyperresponsiveness when
challenged with a dose of acetylcholine. Treatment of mice with antisera to block IL-4 or IgG
production had no effect on diesel induced hyperresponsiveness. Furthermore, histological
examination of the lung tissue clearly showed that when GM-CSF was blocked, tissue injury was
minimal. Otha et al. (1999) could also observe expression of mRNAs in diesel particulate treated
mice for GM-CSF and TNF" which acts to accelerate production of GM-CSF.
Løvik et al. (1997) found that diesel particulate and carbon black when administered together
increase the inflammatory response in lymph nodes that drain the site of deposition. Diesel
particles and carbon black prolong the duration of the inflammatory response. This effect was
synergistic, leading Løvik et al. (1997) to conclude that they had observed an adjuvant effect of
diesel particles on allergic responses. Diesel particles have an irritative effect on epithelial cell
surfaces, so when administered in conjunction with allergenic substances, the facilitate the
inflammatory reaction. Løvik et al. (1997) proposed that chemical substances adsorbed to diesel
exhaust particles could also affect cellular responses, and directly or indirectly elicit mediator
production. They found that the adjuvant activity of diesel particles also lay in the pure carbon
of carbon black.
In vitro studies have been carried out on a number of tissue explants and stably transformed cell
lines to investigate the toxicological responses to diesel particles. Bayram et al. (1998) and Devalia
et al. (1999) have examined the mechanisms associated with the recruitment of inflammatory cells
including eosinophils, neutrophils, lymphocytes and mast cells to air way mucosa following
exposure to diesel particulate. Primary cultures of bronchial epithelial cells were prepared from
bronchial biopsies from 28 human volunteers. Thirteen were non-atopic, non-asthmatic, and the
remainder were atopic and mildly asthmatic. Cultures were treated with increasing doses of diesel
exhaust particulate. Cells of non-asthmatic asymptomatic volunteers showed no dose dependent
release of several cytokines or chemokines including IL-8 or RANTES. RANTES is a potent
chemokine affecting the activity of eosinophils (Bayram et al., 1998). There was evidence for an
increase in GM-CSF in these cultures. Bronchial epithelial explants or atopic and asthmatic
subjects had elevated constitutive levels of all markers of inflammatory response. Exposure to
diesel particles produced significant rises in IL-8, GM-CSF, RANTES and soluble intercellular
adhesion molecule-1 (sICAM-1) (Devalia et al., 1999). These results showed that explants of
bronchial epithelial cells from asthmatic individuals released significantly greater quantities of
specific pro-inflammatory molecules when challenged with diesel particulate matter. These
exposures did not produce evidence of gross cytological damage. Cilliary beat frequency of the
cultured cells was progressively attenuated with increasing concentration of diesel particles in
both asthmatic and asymptomatic individuals (Bayram et al., 1998).
Boland et al. (1999) have examined the effect of diesel exhaust particles from catalyst equipped
or from engines with no catalytic converter on primary cell cultures prepared from nasal epithelial
cells, or on bronchial epithelial cells from a stably transformed cell line. Diesel particles produced
emission.
no obvious signs of cytotoxicity, but there was some evidence of membrane damage based on
release of LDH to the surrounding medium. The uptake of particles by cells changed the relative
transparency and granularity of cells, affording direct observation of engulfed particles. Treated
cells secreted pro-inflammatory cytokines GM-CSF and IL-8. It was proposed by the authors
that the inflammatory response was possibly due to the adsorbed organic compounds including
PAH, PAC and nitro-PAC characteristically found associated with diesel exhaust particles.
A similar report using scanning and transmission electron microscope studies of human
bronchial epithelial cells engulfing diesel particles has been prepared by Steerenberg et al. (1998).
Exposure to diesel particles also induced the production of IL-6 and IL-8, and particles could
be observed adhering the cells as well as inside them.
There has been some interest in the possibility for diesel or other fine particles in urban air to
bind allergens, and thus express the adjuvant activity directly on humans who inhale mixed
particles. Ormstad et al. (1998) have reported on tests to show stable associations of cat dander,
dog hair allergen, birch pollen, and pure D. pteronyssinus (house dust mite allergen). Diesel
particles were found to bind to all allergens tested, and there was some evidence that the cells
responded to the bound allergens.
It has been suggested that free grass pollen allergens from whole pollen or ruptured grass pollen
dispersed in the atmosphere bind to other particles suspended in the atmosphere. Knox et al.
(1997) loaded two different natural allergens from grass onto diesel particles, then demonstrated
their presence by immunofluorescence. Rye grass allergens Lol p1 and Lol p5 are proteins of
approximately 35 Kda molecular weight. The binding to particles is significant because in 1994,
asthmatics living in urban, more polluted areas of London, UK were affected by a thunderstorm.
It was determined that the asthmatic population had responded to finely dispersed grass allergens
that were released by a violent thunderstorm. Pollen and pollen fragments was adsorbed to fine
matter in the urban air, and the particulate surfaces elicited a strong reaction in the lung when
inhaled. The large surface area of the mostly carbon diesel core particles are likely to adsorb all
types of organic material in the correct conditions (Knox et al., 1997).
emission.
The U.S. EPA (2002b) has concluded that it is possible to estimate a reference concentration for
inhalation exposure to diesel exhaust (as measured by diesel particulate matter or DPM). A
reference concentration is defined as that amount of diesel exhaust particulate matter to which
humans may be exposed throughout their lifetime without being likely to experience adverse
noncancer respiratory effects. This exposure level, known as the reference concentration (RfC)
5 :g/m3 of DPM was derived on the basis of dose-response data on inflammatory and
histopathological changes in the lung from rat inhalation studies.
10.3
Cancer and Other Risks from Diesel Emissions
Valberg and Crouch (1999) recently performed a meta-analysis of rat lung tumours from lifetime
inhalation of diesel exhaust. In rats, high concentrations of diesel exhaust particulate (>1 000
:g/m3) inhaled over a lifetime produced excess lung tumours. However, to achieve this level
of exposure, rats suffered massive particle overload, and could not clear the particulate matter.
At doses that do not produce these overload exposures, tumours do not appear. Valberg and
Crouch combined the data for eight different chronic inhalation studies in a meta-analysis, and
were able to establish a threshold of response at somewhere between 200 :g/m3 and 600 :g/m3.
Below a dose rate of 600 :g/m3 exposure for a lifetime, there was no evidence of tumour
production. Although the rat studies were conducted at different laboratories, used somewhat
different protocols and exposures, the results were far more homogeneous that could ever be
expected from an epidemiology study (Valberg and Crouch, 1999).
Using epidemiology data from workers exposed to diesel particulate matter (locomotive and
railroad workers) the California Air Resources Board obtained a unit risk estimate of 300 x 10-6
per :g/m3 (CARB, 1998). This value is inconsistent with the analysis for cancer risk determined
by Valberg and Crouch (1999) who estimated an upper bound unit risk in humans of 9.3 x 10-6
per :g/m3 (using body surface scaling factors) at a lifetime exposure concentration of 478 :g/m3
[CI 155-624 :g/m3]. The upper bound estimate was derived from data collected for 4600
animals (including controls and those exposed below the threshold).
The U.S. EPA has made the assessment that diesel exhaust is “likely to be carcinogenic to
humans by inhalation” and that this hazard applies to environmental exposures (U.S. EPA,
2002b). This conclusion is based on the totality of evidence from human, animal, and other
supporting studies. There is considerable evidence demonstrating an association between diesel
exhaust exposure and increased lung cancer risk among workers in varied occupations where
diesel engines historically have been used. In addition to the human evidence, there is supporting
evidence of diesel particulate matter’s (DPM’s) carcinogenicity and associated DPM organic
compound extracts in rats and mice by non-inhalation routes of exposure. High exposure,
chronic inhalation studies in the rat are considered poor models for assessment of the
carcinogenic potential of diesel exhaust in humans.
Although the available human evidence shows a lung cancer hazard to be a likely consequence
of occupational exposures (generally higher than environmental levels), it is reasonable to
emission.
presume that the hazard extends to environmental exposure levels. There is an incomplete
understanding of the mode of action for diesel exhaust-induced lung cancer that may occur in
humans. Organic extracts of diesel particulate matter contain potent mutagenic compounds.
Therefore, there is the potential for a non-threshold mode of action by components of diesel
exhaust. Even after considering the uncertainties and the assumptions involved, the evidence
for a potential cancer hazard to humans resulting from chronic inhalation exposure to diesel
exhaust is persuasive (U.S. EPA, 2002b). On the other hand, even though there is ample
evidence for a hazard, an assessment of risk can not be made with confidence because of the lack
of a properly defined exposure/dose-response relationship.
emission.
11.0
CARCINOGENIC/GENOTOXIC
URBAN AIR
POTENTIAL
11.1
Epidemiology Studies for diesel emissions
OF
Lung cancer is now the leading cause of cancer mortality among men and women in North
America. PAH and their association with particulate matter have sometimes thought to be a
potential risk factor for lung cancer attributed to urban air pollution sources. The increases in
prevalence of lung cancer have been linked to the use of tobacco products. Tobacco smoke
causes 85-90% of all lung cancer cases, and represents the most preventable risk factor for this
malignancy (Mannino et al., 1998). Epidemiological observations suggest that high levels of
urban air pollution may result in increased risk of lung cancer, sufficient to account for a few
percent (~1-3%) of total lung cancer incidence (Georgiadis and Kyrtopoulos, 1999). In
Denmark, Engholm et al. (1996) found in an epidemiological study that there appeared to be an
threshold for detection of influences from outdoor air on development of lung cancer in the
general population.
Fuel chemistry plays an important part in emission of particle-associated PAHs. Studies have
shown that the solvent-extractable PAHs from diesel particulate originate almost entirely in the
fuel (Williams et al., 1987; Andrews et al., 1998; Hsiao-Hsuan et al., 2000). The PAH molecules
originally present in diesel fuel resist thermal breakdown, so a significant fraction survive the
combustion process and condense onto the diesel particulate matter (DPM). These studies have
been confirmed by other research groups (Crebelli et al., 1995; Tancell et al., 1995). There is a
consensus among these researchers that denovo production of PAH (pyrosynthesis) occurs only
at the highest temperature operating conditions in a diesel engine. Under these conditions, most
of the DPM and other pyrolysis products are ultimately burned before exiting the cylinder.
These results indicate that emissions of PAHs are more a function of the PAH content of the
fuel than of engine technology. For a given refinery and crude oil, diesel fuel PAH correlates
with total aromatic content. PAH have been implicated as one potential contributing
component to the observed toxicity of diesel exhaust (DE), changes in PAH content of diesel
fuel over time, as well as differences between diesel fuels used in different applications, may
influence the hazard observed in exposed populations. Because of the differences in fuel,
engines and engine load, and therefore emissions, a general characterization of diesel particulate
bound PAH is not possible for purposes of differentiating health impact or relative risk in an
epidemiologic study (U.S. EPA, 2002b).
At least 32 PAH have been identified in the exhaust of light- and heavy-duty diesel (LD and
HD) vehicles and (U.S. EPA, 2002b) (see Table 8). In general, among the vehicles tested, PAH
emission rates were higher for LD diesel vehicles compared with HD diesel vehicles. Emission
rates of four representative particle-phase PAHs from HD diesel vehicles, LD diesel vehicles,
and gasoline are shown below. Emission rates for benzo[a]pyrene were higher in diesel
emissions compared with gasoline emissions. These are numbers generated from old
combustion technology, and as newer vehicles enter the fleet, the emissions of PAH per vehicle
emission.
mile should decrease.
Table 8: Emission rates of particle-bound PAH (:g/mile)
from diesel and gasoline enginesA
Heavy duty diesel
Light duty diesel
Gasoline engines
B
C
D
B
E
B
D
Pyrene
71
17.6
36.2
245
66
248
4
Fluroanthene
44
27.2
20.8
213
50
196
3.6
Benzo[a]pyrene
13
<0.1
2.1
13
NA
1
3
Benzo[e]pyrene
10
0.24
4.2
19
NA
1
3.6
A
Adapted from U.S. EPA 2002b
Watson et al., 1998
C
Westerholm et al., 1991
D
Rogge et al., 1993
E
Smith, 1989; Mercedes Benz
NA = data not available
B
The risk imposed by exposure to fine particulate air pollution is obviously much smaller than
the risk associated with cigarette smoking as shown in column three of Table 8 (above). Pope
(Pope et al., 2002) concluded that long-term exposure to particulate air pollution in metropolitan
urban communities across the United States is an important risk factor for cardiopulmonary
mortality including lung cancer. Fine particulate and sulphur oxide-related pollution have been
associated with all-cause mortality, lung cancer, and cardiopulmonary mortality (Pope et al.,
2002). Each incremental increase of 10 :g/m3 in fine PM could be associated with an increase
in all-cause mortality (4%), approximately 6% for cardiopulmonary mortality, and 8% for lung
cancer mortality. Coarse particle fractions did not appear to confirm the associations that were
recognizable based on fine particulate matter concentrations. Pope et al. (2002) have reported
that combustion-related fine particulate matter is an important risk factor for lung cancer,
especially among groups with low socioeconomic status.
There are multiple etiological factors that relate lung cancer mortality and exposure to airborne
PM or chemical air toxics. The levels of health effects measured by lung cancer mortality after
chronic exposure to ambient air pollution has been the subject of a recent review (Beeson et al.,
1998). Lung cancer incidence in males was associated with PM10 exceedance frequencies of 40,
50, 60, 80, and 100 :g/m3 with the regression estimates increasing as the cutoff increased. For
females, the relative risks of lung cancer incidence were all above 1.0 for each of the PM10
thresholds investigated. Results were determined from an epidemiology study on a population
(AHSMOG) living in California, and exposed to ambient air quality for the period 1977-1992.
emission.
11.2
Extrapolation issues for lung cancer from rodent studies.
A large number of studies have examined the effect of various forms of particulate matter on
pulmonary tissues. Rittinghausen et al. (1997) have recently reviewed a series of rat inhalation
studies for the effects of PM or chemical exposures to coal oven gas, coal tar/pitch
condensation aerosol, or diesel engine exhaust. Results confirmed that chronic inhalation
studies of PM show that a high burden of particles leads to cystic keratinizing epithelioma, a benign
neoplasm. The combination of particles with instillation of known carcinogens such as B[a]P,
crocidolite or dibenz[a,h]anthracene leads to production of invasive cystic keratinizing squamous
cell carcinomas. It is suspected that the benign neoplasm is only a stage on the way to the fully
malignant carcinoma (Rittinghausen et al.,1997).
The epidemiological evidence for an association between diesel particles and elevated lung
cancer incidence is variable. There is a discrepancy between results of chronic inhalation, and
instillation studies in rodents and the poorly defined human epidemiological evidence that
suggests rodent inhalation studies may not be relevant for humans. There is a clear difference
in the location of particle deposition in lungs of rodents and primates. Rats show a differential
deposition of particles in the intra-alveolar spaces, while evidence for particle deposition in
alveoli of the monkey exposed to carbon black for years is quite different. In Primates, particles
were localized in peribronchial or perivascular interstitial tissues (Rittinghausen et al.,1997).
Thus, the evidence of chronic inflammatory response induced by alveolar macrophages in rats
exposed to particles is absent in monkeys. This could be due to a lower uptake of particles by
alveolar macrophages, and hence a decreased inflammatory response. In rats there is a strong
participation of polymorphonuclear leukocytes (PMN) associated with reactive oxygen species
(ROS). This response is important because of its mitogenic, genotoxic and cytotoxic
associations and consequences. In humans known to have chronic exposure to high levels of
PM, relatively few PMN have been observed in lung lavages. The result of these studies reveal
that PMN do not seem to be a significant part of pulmonary response to PM in humans
(Rittinghausen et al.,1997).
Since (1) particle-induced chronic inflammation of lung tissue eventually leading
to neoplastic changes is considered fundamental to lung tumour induction in
rats, and (2) rats are different to man and other species in many respects of their
particle-induced inflammatory response, it is considered not possible to make direct
extrapolation of results in rats to humans (Rittinghausen et al.,1997).
emission.
12.0
REFERENCES
Abbey DE, Nishino N, McDonnell WF, Burchette RJ, Knutsen SF, Beeson WL, Yang JX. 1999.
Long-term inhalable particles and other air pollutants related to mortality in nonsmokers. Am.
J. Respir. Crit. Care Med. 159: 373-382.
Abbey DE, Hwang BL, Burchette RJ, Vancuren T, Mills PK. 1995. Estimated long-term
ambient concentrations of PM10 and development of respiratory symptoms in a nonsmoking
population. Arch Environ Health 50:139-152. (Cited in Künzli et al., 2001)
Abbey DE, Burchette RJ, Knutsen SF, McDonnell WF, Lebowitz MD, Enright PL. 1998.
Long-term particulate and other air pollutants and lung function in nonsmokers. Am J Respir
Crit Care Med, 158: 289-298.
Ackerman-Liebrich U, Leunenberger P, Schwartz J, Schindler C, Monn C, Bolognini G,
Bongard JP, Brändli O, Domenighetti G, Elsasser S, Grize L, Karrer W, Keller R, KellerWossidilo H, Künzli N, Martin BW, Medici TC, Perruchoud AP, Schöni MH, Tschopp JM,
Villiger B, Wüthrich B, Zellweger JP, Zemp E, and the Study on Air Pollution and Lung
Diseases in Adults (SAPALDIA) Team. 1997. Lung function and long term exposure to air
pollutants in Switzerland. Am J Respir Crit Care Med 155: 122-129.
Ackerman-Liebrich U, Leunenberger P, Schwartz J, Schindler C, Monn C, Bolognini G,
Bongard JP, Brändli O, Domenighetti G, Elsasser S, Grize L, Karrer W, Keller R, KellerWossidilo H, Künzli N, Martin BW, Medici TC, Perruchoud AP, Schöni MH, Tschopp JM,
Villiger B, Wüthrich B, Zellweger JP, Zemp E, and the Study on Air Pollution and Lung
Diseases in Adults (SAPALDIA) Team. 1997. Lung function and long-term exposure to air
pollutants in Switzerland. Am J Respir Crit Care Med 155: 122-129.
Adamson IY, Prieditis H, Hedgecock C, Vincent R. 2000. Zinc is the toxic factor in the lung
response to an atmospheric particulate sample. Toxicol Appl Pharmacol 166(2): 111-119.
Adamson I.Y.R, Vincent R, Bjarnason S.G. 1999a. Cell injury and interstitial inflammation in
rat lung after inhalation of ozone and urban particulates. Am J Respir Cell Mol Biol 20:10671072.
Adamson I.Y.R, Prieditis H, Vincent R. 1999b. Pulmonary toxicity of an atmospheric
particulate sample is due to the soluble fraction. Toxicol Appl Pharmacol 157:43-50.
Allen PD. 1987. Residential population exposure to ambient benzene in California. California
Air Resources Board. ARB/TS-87-001 (cited in ATSDR, 1997).
Allred EN, Bleecker ER, Chaitman BR, Dahms TE, Gottlieb SO, Hackney JD, Pagano M,
Selvester RH, Walden SM, Warren J. 1991. Effects of carbon monoxide on myocardial
ischemia. Environ Health Perspect 91: 89-132.
emission.
Alm S, Jantunen MJ, Vartiainen M. 1999. Urban commuter exposure to particle matter and
carbon monoxide inside an automobile. J Expos Anal Epidemiol, 9: 237-244.
Anderson HR, Ponce de Leon A, Bland JM, Bower JS, Emberlin J, Strachan DP. 1998. Air
pollution, pollen, and daily admissions for asthma in London. Thorax, 53: 842-848.
Andrews GE, Ishaq RB, Farrar-Khan JR, et al. 1998. The influence of speciated diesel fuel
composition on speciated particulate SOF emissions. SAE Technical Paper Ser. No. 980527 (and
references therein). (Cited in U.S. EPA, 2002b)
Atkinson RW, Anderson HR, Strachan DP, Bland JM, Bremmer SA, Ponce de Leon A. 1999.
Short term associations between outdoor air pollution and visits to accident and emergency
departments in London for respiratory complaints. 1999. Eur Respir, 13: 257-265.
ATSDR. 1997. Toxicological Profile for Benzene (Update) U.S. Dept Health & Human Services,
Public Health Service Agency for Toxic Substances and Disease Registry. September, 1997.
Basha MA, Gross KB, Gwizdala CJ, Haidar AH, Popovich J;. 1994. Bronchoalveolar lavage
neutrophilia in asthmatic and healthy volunteers after controlled exposure to ozone an filtered
purified air. Chest, 106: 1757-1765.
Bates DV. 2000. Lines that connect: Assessing the causality inference in the case of particulate
pollution. Environ Health Perspect, 108: 91-92.
Bayram H, Devalia JL, Khair OA, Abdelaziz MM, Sapsford RJ, Sagai M, Davies RJ. 1998.
Comparison of ciliary activity and inflammatory mediator release from bronchial epithelial cells
of nonatopic nonasthmatic subjects and atopic asthmatic patients and the effect of diesel exhaust
particles in vitro. J Allergy Clin Immunol, 102:771-782.
Beeson LW, Abbey DE, Knutsen SF. 1998. Long-term concentrations of ambient air pollutants
and incident lung cancer in California adults: Results from the AHSMOG study. Environ
Health Perspect, 106: 813-823.
Bergh M, Karlberg A. 1999. Sensitizing potential of acetaldehyde and formaldehyde using a
modified cumulative contact enhancement test (CCET. Contact Dermatitis, 40: 139-145.
Bobak M. 2000. Outdoor air pollution, low birth weight, and prematurity. Environ Health
Perspect, 108: 173-176.
Boezen M, Schouten J, Rijcken B, Vonk J, Gerritsen J, van der Zee S, Hoek G, Brunekreef B,
Postma D. 1998. Peak expiratory flow variability, bronchial responsiveness, and susceptibility
to ambient air pollution in adults. Am J Respir Crit Care Med 158: 1848-1854.
emission.
Boezen HM, van der Zee S, Postma D, Vonk J, Gerritsen J, Hoek G, Brunekreef B, Rijcken B,
Schouten J. 1999. Effects of ambient air pollution on upper and lower respiratory symptoms
and peak expiratory flow in children. The Lancet 353: 874-878.
Boland S, Baeza-Squiban A, Fournier T, Houcine O, Gendron MC, Chevrier M, Jouvenot G,
Coste A, Aubier M, Marano F. 1999. Diesel exhaust particles are taken up by human airway
epithelial cells in vitro and alter cytokine production. Am J Physiol, 276 (4 Pt 1) pL604-613.
Borchers MT, Carty MP, Leikauf GD. 1999. Regulation of human airway mucins by acrolein
and inflammatory mediators. Am J Physiol, 276(4 Pt 1): L549-555.
Boudet, C., Dechenaux, J., Balducci, F., Masclet, P., Zmirou, D. 1998. Total personal exposure
to fine particulate matter in an urban adult population: The role of ambient air and transport.
International Journal of Vehicle Design, 20: 30-38
Brauer M, Hirtle RD, Hall AC, Yip TR. 1999. Monitoring oersonal fine particle exposure with
a particle counter. J Expos Anal Epidemiol, 9: 228-236.
Brauer M, Hoek G, Van Vliet P, Meliefste K, Fischer PH, Wijga A, Koopman LP, Neijens
HJ, Gerritsen J, Kerkhof M, Heinrich J, Bellander T, Brunekreef B. 2002. Air pollution from
traffic and the development of respiratory infections and asthmatic and allergic symptoms
in children. Am J Respir Crit Care Med 166:1092-1098.
Braun-Fahrländer C, Vuille JC, Sennhauser FH, Neu U, Künzel T, Grize L, Gassner M,
Minder C, Schindler C, Varonier HS, Wüthrich B, and the SCARPOL Team. 1997.
Respiratory health and long-term exposure to air pollutants in Swiss schoolchildren. Am J
Respir Crit Care Med 155: 1042-1049.
Brook JR, Dann TF, Burnett RT. 1997. The relationship among TSP, PM10, PM2.5 and
inorganic constituents of atmospheric particulate matter at multiple Canadian locations. J
Air Waste Manage Assoc, 46: 2-19.
Brunekreef B, Hoek G. 2000b. Invited commentary - Beyond the body count: air pollution
and death. Comment on: Am J Epidemiol 151: 440-448. Am J Epidemiol 151: 449-451.
Brunekreef B, Hoek G, Fischer P, Spieksma FT. 2000a. Relation between airborne pollen
concentrations and daily cardiovascular and respiratory-disease mortality [letter]. Lancet
355 (9214):1517-1518.
Brunekreef B. 1997. Air pollution and life expectancy: is there a relation? Occup Environ
Med 54: 781-784.
Burnett R.T, Cakmak S, Brook J.R and Krewski D. 1997. The role of particulate size and
emission.
chemistry in the association between summertime ambient air pollution and hospitalization
for cardiorespiratory diseases. Environ Health Perspect 105(6):614-620.
Burnett RT, Cakmak S, Raizenne ME, Stieb D, Vincent R, Krewski D, Brook JR, Philips O,
Ozkaynak H. 1998. The association between ambient carbon monoxide and daily mortality
in Toronto, Canada. J Air Waste Manag Assoc 48:689-700.
Burnett RT, Smith-Doiron M, Cakmak S, Brook JR. 1999. Effects of particulate and gaseous
air pollution on cardiorespiratory hospitalizations. Archiv Eniron Health, 54: 130-138.
Burnett RT, Brook J, Dann T, Delocla C, Philips O, Cakmak S, Vincent R, Goldberg MS,
Krewski D. 2000. Association between particulate- and gas-phase components of urban air
pollution and daily mortality in eight Canadian cities. Inhalation Toxicol 12(Suppl 4): 15-39.
Burnett RT, Dales R.E, Raizenne M.E, Krewski D, Summers P.W, Roberts G.R, RaddYoung M, Dann T, Brook J. 1994. Effects of low ambient levels of ozone and sulfates on
the frequency of respiratory admissions to Ontario hospitals. Environ Res 65:172-194.
Burnett, R.T. 1996. Assessing the use of administrative health and environmental
information to link exacerbation of cardio-respiratory diseases to ambient air pollution in
Canada. Clean Air Strategic Alliance Acidifying Emissions Symposium ’96.
Burton RM, Suh HH, Koutrakis P. 1996. Spacial variation in particulate concentrations
within metropolitan Philadelphia. Environ Sci Technol, 30: 400-407.
Calderón-Garcidueñas L, Rodríguez-Alcaraz A, Villarreal-Calderón A, Lyght O, Janszen,
Morgan KT. 1998. Nasal epithelium as a sentinel for airborne environmental pollutants.
Toxicol Sci, 46: 352-364.
Campen MJ, Costa DL, Watkinson WP. 2000. Cardiac and thermoregulatory toxicity of
ROFA in cardiopulmonary compromised rats. Inhalation Toxicol 12(Suppl 2): 7-22.
Campen MJ, Nolan JP, Schladweiler MCJ, Kodavanti UP, Evansky PA, Costa DL,
Watkinson WP. 2001. Cardiovascular and thermoregulatory effects of inhaled PMassociated transition metals: A potential interaction between nickel and vanadium sulfate.
Toxicol Sci 64: 243-252.
emission.
CARB. 1998. Proposed Identification Of Diesel Exhaust As A Toxic Air Contaminant.
Office of Environmental Health Hazard Assessment, California Environmental Protection
Agency Air Resources Board.
Castillejos M, Borja-Aburto VH, Dockery DW, Gold DR, Loomis D. 2000. Airborne coarse
particles and mortality. Inhalation Toxicol 12(Suppl 1): 67-72.
Checkoway H, Levy D, Sheppard L, Kaufman J, Koenig J, Siscovick D. 2000. A CaseCrossover Analysis of Fine Particulate Matter Air Pollution and Out-of-Hospital Sudden
Cardiac Arrest. Research Report 99. Health Effects Institute, Cambridge MA.
Chen P-C, Lai Y-M, Chan C-C, Hwang J-S, Yang C-Y, Wang J-D (1999. Short-term effect
of ozone on the pulmonary function of children in primary school. Environ Health Perspect,
107: 921-925.
Chow JC, Watson JG, Green MC, Lowenthal DH, DuBois DW, Kohl SD, Egami RT, Gilles
J, Rogers CF, Frazier CA. 1999. Middle- and neighbourhood-scale variations of PM10 source
contributions in Las Vegas, Nevada. J Air & Waste Manage Assoc, 49: 641-654.
Chung FL, Zhang L, Ocando JE, Nath RG. 1999. Role of 1,N2-propanodeoxyguanosine
adducts as endogenous DNA lesions in rodents and humans. IARC Sci Publ, 150: 45-54
Claiborn, C.S., Finn, D., Larson, T.V., Koenig, J.Q. 2000. Windblown dust contributes to
high PM2.5 concentrations. J Air Waste Manag Assoc 50: 1440-1445.
Clarke RW, Coull B, Reinisch U, Catalano P, Killingsworth CR, Koutrakis P, Kavouras I,
Krishna Murthy GG, Lawrence J, Lovett E, Wolfson JM, Verrier RL, Godleski JJ. 2000.
Inhaled concentrated ambient particles are associated with hematologic and bronchoalveolar
lavage changes in canines. Environ Health Perspect 108: 1179-1187.
Crebelli R, Conti L, Crochi B, Carere A, Bertoli C, Del Giacomo N. 1995. The effect of fuel
composition on the mutagenicity of diesel engine exhaust. Mutat Res 346:167-172. (Cited
in U.S. EPA, 2002b)
Crosignani P, Borgini A, Cadum E, Mirabelli D, Porro E. 2002. New directions: air
pollution - how many victims? Atmos Environment 36: 4705-4706.
Dejmek J, Selevan SG, Benes I, Solansky I, Šrám RJ. 1999. Fetal growth and maternal
exposure to particulate matter during pregnancy. Environ Health Perspect 107: 475-480.
Devalia JL, Bayram H, Abdelaziz MM, Sapsford RJ, Davies RJ. 1999. Differences between
cytokine release from bronchial epithelial cells of asthmatic patients and non-asthmatic
emission.
subjects: effect of exposure to diesel exhaust particles. Int Arch Allergy Immunol, 118:
437-439.
Dockery D.W. 1993. Epidemiologic study design for investigating respiratory health effects
of complex air pollution mixtures. Environ Health Perspect Suppl 101(S4):187-191.
Dockery DW, Speizer FE, Stram DO, Ware JH, Spengler JD, Ferris BG. 1989. Effects of
inhalable particles on respiratory health of children. Am Rev Respir Dis, 139: 587-594.
Dockery DW, Schwartz J, Spengler JD. 1992. Air pollution and daily mortality: associations
with particulates and acid aerosols. Environ Res 59:362-373.
Dockery DW, Pope CA 3rd, Xu X, Spengler JD, Ware JH, Fay ME, Ferris BG Jr, Speizer
FE. 1993b. An association between air pollution and mortality in six U.S. cities. N Engl J
Med. 329: 1753-1759.
Dockery DW, Pope CA III. 1994. Acute respiratory effects of particulate air pollution. Annu
Rev Public Health 15: 107-132.
Dockery D.W, Cunningham J, Damokosh A.I, Neas L.M, Spengler J.D, Koutrakis P, Ware
J.H, Raizenne M. and Speizer F.E. 1996. Health effects of acid aerosols on North American
children: Respiratory symptoms. Environ Health Perspect 104(5):500-505.
Dockery DW, Pope III CA, Xu X, Spengler JD, Ware JH, Fay ME, Ferris BG Jr, Speizer FE.
1993b. An association between air pollution and mortality in six U.S. cities. N Engl J Med.
329: 1753-1759.
Dominici F, McDermott A, Zeger SL, Samet JM. 2002. On the use of generalized additive
models in time-series studies of air pollution and health. Am J Epidemiol 156: 193-203.
Donaldson K, Brown DM, Mitchell C, Dineva M, Beswick PH, Gilmour P, MacNee W.
1997. Free radical activity of PM10: Ironmediated generation of hydroxyl radicals. Environ
Health Perspect 105(Suppl 5): 1285-1289.
Dye JA, Adler KB, Richards JH, Dreher KL. 1997. Epithelial injury induced by exposure
to residual oil fly-ash particles: role of reactive oxygen species? Am J Respir Cell Mol Biol,
17(5):625-33
Engholm G, Palmgren F, Lynge E. 1996. Lung cancer, smoking, and environment: a cohort
study of the Danish Population. BMJ, 312: 1259-1263.
English P, Neutra R, Scalf R, Sullivan M, Waller L, Zhu L. 1999. Examining associations
emission.
between childhood asthma and traffic flow using a geographic information system. Environ
Health Perspect 107: 761-767.
Environment Canada. 1999. Ambient Particulate Matter An Overview. From National
Ambient Air Quality Objectives for Particulate Matter: Part 1: Science Assessment
Document. 1999. by the Federal Provincial Working Group on Air Quality Objectives and
Guidelines. Section 13 (pp1-5).
Fairley D. 1999. Daily mortality and air pollution in Santa Clara County, California: 19891996. Environ Health Perspect, 107: 637-641.
Galizia A, Kinney PL. 1999. Long-term residence in areas of high ozone: associations with
respiratory health in a nationwide sample of nonsmoking young adults. Environ Health
Perspect, 107: 675-679.
Gauderman WJ, Gilliland GF, Vora H, Avol E, Stram D, McConnell R, Thomas D, Lurmann
F, Margolis HG, Rappaport EB, Berhane K, Peters JM. 2002. Association between air
pollution and lung function growth in southern California children: results from a second
cohort. Am J Respir Crit Care Med. 166: 76-84.
Georgiadis P, Kyrtopoulous SA. 1999. Molecular epidemiological approached to the study
of the genotoxic effects of urban air pollution. Mutat Res, 428: 91-98.
Ghio AJ, Kim C, Devlin RB. 2000. Concentrated ambient air particles induce mild
pulmonary inflammation in healthy human volunteers. Am J Respir Crit Care Med 162: 981988.
Godleski JJ, Verrier RL, Koutrakis P, Catalano P, Coull B, Reinisch U, Lovett EG, Lawrence
J, Murphy GGK, Wolfson JM, Clarke RW, Nearing BD, Killingsworth C. 2000.
Mechanisms of morbidity and mortality from exposure to ambient air particles. Health
Effects Institute Research Report Number 91, Health Effects Institute, Cambridge MA.
Gold DR, Litonjua A, Schwartz J, Lovett E, Larson A, Nearing B, Allen G, Verrier M,
Cherry R, Verrier R. 2000. Ambient pollution and heart rate variability. Circulation 101:
1267-1273.
emission.
Gold DR, Damokosh AI, Pope CA III, Dockery DW, McDonnell WF, Serrano P, Retama A,
Castittejos. 1999. Particulate and ozone pollutant effects on the respiratory function of
children in Southwest Mexico City. Epidemiology 10: 8-16.
Goldberg MS, Bailar JC III, Burnett RT, Brook JR, Tamblyn R, Bonvalot Y, Ernst P, Flegel
KM, Singh RK, Valois M-F. 2000. Identifying Subgroups of the General Population That
May Be Susceptible to Short-Term Increases in Particulate Air Pollution: A Time-Series
Study in Montreal, Quebec. Research Report 97. Health Effects Institute, Cambridge MA.
Gordian M.E, Ozkaynak H, Xue J, Morris S.S. and Spengler J.D. 1996. Particulate air
pollution and respiratory disease in Anchorage, Alaska. Environ Health Perspect 104(3):290297.
Gordon T, Nadziejko C, Schlesinger R, Chen LC. 1998. Pulmonary and cardiovascular
effects of acute exposure to concentrated ambient particles in rats. Toxicol Lett 96-97: 285288.
Guo YL, Lin Y-C, Sung F-C, Huang S-L, Ko Y-C, Lai J-S, Su H-J, Shaw C-K, Lin R-S,
Dockery DW. 1999. Climate, traffic-related air pollutants, and asthma prevalence in middleschool children in Tiawan. Environ Health Perspect, 107: 1001-1006
Gwynn CR, Burnett RT, Thurston GD. 2000. A time-series analysis of acidic particulate
matter and daily mortality and morbidity in the Buffalo, New York, region. Environ Health
Perspect, 108: 125-133.
Habbick BF, Pizzichini MMM, Taylor B, Rennie D, Senthilselvan A, Sears MR. 1999.
Prevalence of asthma, rhinitis and eczema among children in 2 Canadian cities: the
International Study of Asthma and Allergies in Childhood. CMAJ;160:1824-1828.
HEI 2000. Samet JM, Zeger SL, Dominici F, Curriero F, Coursac I, Dockery DW, Schwartz
J, Zanobetti A. 2000. The National Morbidity, Mortality, and Air Pollution Study Part II:
Morbidity, Mortality, and Air Pollution in the United States. Health Effects Institute, Report
Number 94, Part II. Cambridge MA www.healtheffects.org
HEI. 1995. Health Effects Institute. Introduction. Diesel Exhaust: A critical analysis of
emissions, exposure and health effects. A special Report on the Institute’s Working Group,
Nauss KM and the HEI Diesel Working Group. Pp 13-61.
HEI. 2002. Understanding The Health Effects of Components of The Particulate Matter
Mix: Progress And Next Steps. Perspectives Insights from HEI’s research programs. Health
Effects Institute, Cambridge MA. April 2002. http://www.healtheffects.org
emission.
HEI. 1999. Health Effects Institute: Program Summary, Research on Air Toxics. May 1999
Heinrich J, Hoelscher B, Wjst M, Cyrys J, Wichmann HE. 1999. Respiratory diseases and
allergies in two polluted areas in East Germany. Environ Health Perspect 107: 52-62.
Hoek G, Brunekreef B. 1993. Acute effects of a winter air pollution episode on pulmonary
function and respiratory symptoms of children. Arch. Environ. Health 48: 328-335.
Hoek G, Brunekreef B, Verhoeff A, van Wijnen J, Fischer P. 2000. Daily mortality and air
pollution in The Netherlands. J Air Waste Manag Assoc 50: 1380-1389.
Hoek G, Schwartz JD, Groot B, Eilers P. 1997. Effects of ambient particulate matter and
ozone on daily mortality in Rotterdam, the Netherlands. Arch Environ Health 52:455-463.
Horak F Jr, Studnicka M, Gartner C, Spengler JD, Tauber E, Urbanek R, Veiter A, Frischer T.
2002. Particulate matter and lung function growth in children: a 3-yr follow-up study in Austrian
schoolchildren. Eur Respir J. 19: 838-845.
Hsiao-Hsuan M, Wen-Jhy L, Chung-Ban C, et al. 2000. Effect of fuel aromatic content on
PAH emission from a heavy-duty diesel engine. Chemosphere 41:1783-1790. (Cited in U.S.
EPA, 2002b)
Ibald-Mulli A, Stieber J, Wichmann H-E, Koenig W, Peters A. 2001. Effects of air pollution
on blood pressure: A populationbased approach. Am J Public Health 91: 571-577.
Jammes Y, Delpierre S, Delvolgo MJ, Humbert-Téna C, Burnet H. 1998. Long-term
exposure of adults to air pollution is associated with increased airway obstruction and higher
prevalence of bronchial hyperresponsiveness. Archives of Environmental Health 53: 372377.
Janssen, NAH., Schwartz J, Zanobetti A, Suh HH. 2002. Air Conditioning and SourceSpecific Particles as Modifiers of the Effect of PM10 on Hospital Admissions for Heart and
Lung Disease. Environmental Health Persepctives, 110: 43-49.
Janssen, NAH, Hoek G, Harssema H, Brunekreef B. 1999. Personal exposure to fine
particles in children correlates closely with ambient fine particles. Archives of
Environmental Health, 54(2): 95-101.
emission.
Janssen NAH, Hoek G, Harssema H, Brunekreef B. 1998a. Personal sampling of airborne
particles: Method performance and data quality. Journal of Exposure Analysis and
Environmental Epidemiology, 8(1): 37-49.
Janssen NAH, Hoek G, Brunekreef B, Harssema H, Mensink I, Zuidhof A. 1998b. Personal
sampling of particles in adults: Relation among personal, indoor, and outdoor air
concentrations. American Journal of Epidemiology, 147(6): 537-547.
Jedrychowski W, Flak E, Mróz E. 1999. The adverse effect of low levels of ambient air
pollutants on lung function growth in preadolescent children. Environ Health Perspect 107:
669-674.
Katsouyanni K, Karakatsani A, Messari I, Touloumi G, Hatzakis A, Kalandidi A,
Trichopoulos D. 1990. Air pollution and causespecific mortality in Athens. J Epidemiol
Commun Health 44: 321-324.
Kelsall JE, Samet JM, Zeger SL Xu J. 1997. Air pollution and mortality in Philadelphia,
1974-1988. Am J Epidemiol, 146: 750-762.
Knox RB, Suphioglu C, Taylor P, Desai R, Watson HC, Peng JL, Bursill LA. 1997. Major
grass pollen allergen Lol p 1 binds to diesel exhaust particles: implications for asthma and
air pollution. Clin Exp Allergy, 27: 246-251.
Kodavanti UP, Schladweiler MC, Ledbetter AD, Watkinson WP, Campen MJ, Winsett DW,
Richards JR, Crissman KM, Hatch GE, Costa DL. 2000. The spontaneously hypertensive
rat as a model of human cardiovascular disease: Evidence of exacerbated cardiopulmonary
injury and oxidative stress from inhaled emission particulate matter. Toxicol Appl Pharmacol
164: 250-263.
Kodavanti UP, Hauser R, Christiani DC, Meng ZH, McGee J, Ledbetter A, Richards J, Costa
DL. 1998. Pulmonary responses to Oil Fly Ash Particles in the rat differ by virtue of their
specific soluble metals. Toxicological Sci, 43: 204-212.
Koenig JQ, Larson TV, Hanley QS, Rebolledo V, Dumler K, Checkoway H et al.,. 1993.
Pulmonary function changes in children associated with fine particulate matter. Environ Res,
63: 26-38.
emission.
Koenig JQ. 1999. Air pollution and asthma. J Allergy Clin Immunol, 104 (4 pt 1): 717-722.
Künzli N, Medina S, Kaiser R, Quénel P, Horak Jr. F, Studnicka M. 2001. Assessment of
deaths attributable to air pollution: should we use risk estimates based on time series or on
cohort studies? Am J Epidemiol 153: 1050-1055.
Künzli N, Kaiser R, Medina S, Studnicka M, Chanel O, Filliger P, Horak M, PuybonnieuxTexier V, Quénel P, Schneider J, Seethaler R, Vergnaud J-C, Sommer H. 2000. Publichealth impact of outdoor and traffic-related air pollution: a European assessment. Lancet
356: 795-801.
Künzli N. 2002. The public health relevance of air pollution abatement. Eur Respir J 20:
198-209.
Laden F, Neas LM, Dockery DW, Schwartz J. 2000. Association of Fine Particulate Matter
from Different Sources with Daily Mortality in Six U.S. Cities. Environmental Health
Perspectives. 108: 941-947.
Lee J-T, Shin D, Chung Y. 1999. Air pollution and daily mortality in Seoul and Ulsan,
Korea. Environ Health Perspect, 107: 149-154.
Lee JT, Shy CM. 1999. Respiratory function as measured by peak expiratory flow rate and
PM10: six communities study. J Expo Anal Environ Epidemiol, 9: 293-299.
Levy JI, Hammitt JK, Spengler JD. 2000. Estimating the mortality impacts of particulate
matter: what can be learned from between-study variability? Environ Health Perspect, 108:
109-117.
Liao D, Creason J, Shy C, Williams R, Watts R, Zweidinger R. 1999. Daily variation of
particulate air pollution and poor cardiac autonomic control in the elderly. Environ Health
Perspect, 107: 521-525.
Lioy PJ, Wainman T, Zhang J. 1999. Typical household vacuum cleaners: The collection
efficiency and emissions characteristics for fine particles. J. Air & Waste Manage. Assoc.,
49: 200-206.
Lipfert FW . 1993. A critical review of studies of the association between demands for
hospital services and air pollution. Environ Health Perspect Suppl 101(Suppl. 2):229-268.
Lipfert FW, Wyzga RE. 1996. The effects of exposure error on environmental
Epidemiology. Presented at the 2nd Colloquium on Particulate Air Pollution and Health,
Park City, UT. May 1996. Also submitted to Water, Soil and Air Pollution.
emission.
Lipfert FW;. 1997. Air pollution and human health: Perspectives for the ‘90s and beyond.
Risk Analysis 17(2):137-146.
Lipfert FW, Wyzga RE. 1995. Uncertainties in identifying responsible pollutants in
observational epidemiology studies. Inhalation Toxicol 7:671-689.
Lippmann M, Ito K, Nádas A, Burnett RT. 2000. Association of Particulate Matter
Components with Daily Mortality and Morbidity in Urban Populations. Research Report 95.
Health Effects Institute, Cambridge MA.
Loomis D, Castillejos M, Gold DR, McDonnell W, Borja- Aburto VH. 1999. Air pollution
and infant mortality in Mexico City. Epidemiology 10: 118-123.
Løvik M, Høgseth AK, Gaarder PI, Hagemann R, Eide I. 1997. Diesel exhaust particles and
carbon black have adjuvant activity on the local lymph node response and systemic IgE
production to ovalbumin. Toxicology, 121: 165-178.
Mannino DM, Ford E, Giovino GA, Thun M. 1998. Lung cancer deaths in the United States
from 1979 to 1992: An analysis using multiple-cause mortality data. Int. J Epidemiol, 27:
159-166.
Maynard AD, Maynard RL. 2002. A derived association between ambient aerosol surface
area and excess mortality using historic time series data. Atmos Environ 36: 5561-5567.
McClellan RO and Miller FJ. 1997. An Overview of EPA's Proposed Revision of the
Particulate Matter Standard. CIIT Activities. Vol.17, No.4), Chemical Industry Institute of
Toxicology.
McConnell R, Berhane K, Gilliland F, London SJ, Vora H, Avol E, Gauderman J, Margolis
HG, Lurmann F, Thomas DC, Peters JM. 1999. Air pollution and bronchitic symptoms in
Southern California children with asthma. Environ Health Perspect 107: 757-760.
McMichael AJ, Anderson HR, Brunekreef B, Cohen AJ. 1998. Inappropriate use of daily
mortality analyses to estimate longer-term mortality effects of air pollution. Int J Epidemiol
27: 450-453.
McMichael AJ, Anderson HR, Brunekreef B, Cohen AJ. 1998. Inappropriate use of daily
mortality analyses to estimate longer-term mortality effects of air pollution. Int J Epidemiol
27: 450-453.
emission.
Miguel AG, Cass GR, Glovsky MM, Weiss J. 1999. Allergens in paved road dust and
airborne particles. Environ Sci Technol, 33: 4159-4168.
Miyabara Y, Yanagisawa R, Shimojo N, Takano H, Lim HB, Ichinose T, Sagai M. 1998b.
Murine strain differences in airway inflammation caused by diesel exhaust particles. Eur
Respir J, 11: 291-298.
Miyabara Y, Ichinose T, Takano H, Lim HB, Sagai M. 1998a. Effects of diesel exhaust on
allergic airway inflammation in mice. J Allergy Clin Immunol, 102:805-812.
MOE. 1999. A Compendium of Current Knowledge on Fine Particulate Matter in Ontario.
Prepared by the Ontario Ministry for the Environment for the Ontario Smog Plan Steering
Committee. As released to the CHRESTech/Neram Expert Panel for Review. March 1999
Monn C, Becker S. 1999. Cytotoxicity and induction of proinflammatory cytokines from
human monocytes exposed to fine (PM2.5) and coarse particles (PM10-2.5) in outdoor and
indoor air.
Toxicol Appl Pharmacol;155(3):245-52
Mortimer KM, Redline S, Kattan M, Wright EC, Kercsmar CM. 2001. Are peak flow and
symptom measures good predictors of asthma hospitalizations and unscheduled visits?
Pediatric Pulmonology 32: 190-197.
Moyer CF, Kodavanti UP, Haseman JK, Costa DL, Nyska A. 2002. Systemic vascular
disease in male B6C3F1 mice exposed to particulate matter by inhalation: Studies conducted
by the National Toxicology Program. Toxicologic Pathology 30: 427-434.
Nel AE, Diaz-Sanchez D, Ng D, Hiura T, Saxon A. 1998. Enhancement of allergic
inflammation by the interaction between diesel exhaust particles and the immune system.
J Allergy Clin Immunol, 102 (4 Pt 1): 539-554.
Nel AE, Diaz-Sanchez D, Li N. 2001. The role of particulate pollutants in pulmonary
inflammation and asthma: Evidence for the involvement of organic chemicals and oxidative
stress. Curr Opin Pulm Med 7: 20-26.
Nemmar A, Hoet PH, Vanquickenborne B, Dinsdale D, Thomeer M, Hoylaerts MF,
Vanbilloen H, Mortelmans L, Nemery B. 2002. Passage of inhaled particles into the blood
circulation in humans. Circulation 105: 411-414.
Nemmar A, Vanbilloen H, Hoylaerts MF, Hoet PH, Verbruggen A, Nemery B. 2001.
Passage of intratracheally instilled ultrafine particles from the lung into the systemic
circulation in hamster. Am J Respir Crit Care Med 164: 1665-1668.
emission.
Nordenhäll C, Pourazar J, Leden M-C, Levin J-O, Sandström T, Ädelroth E. 2001. Diesel
exhaust enhances airway responsiveness in asthmatic subjects. Eur Respir J 17: 909-915.
Norris G, Young Pong SN, Koenig JQ, Larson TV, Sheppard L, Stout JW. 1999. An
association between fine particles and asthma emergency department visits for children in
Seattle. Environ Health Perspect, 107: 489-493.
Ohta K, Yamashita N, Tajima M, Miyasaka T, Nakano J, Nakajima M, Ishii A, Horiuchi T,
Mano K, Miyamoto T. 1999. Diesel exhaust particulate induces airway hyperresponsiveness
in a murine model: essential role of GM-CSF. J Allergy Clin Immunol, 104:1024-1030.
Oldham J, Mannix RC, Phalen RF. 1997. Deposition of monodisperse particles in hollow
models representing adult and child-size tracheobronchial airways. Health Physics, 72: 827834.
OMA. 1998. Health Effects of Ground-Level Ozone, Acid Aerosols and Particulate Matter.
Ontario Medical Association Position Paper, May, 1998.
Ormstad H, Johansen BV, Gaarder PI. 1998. Airborne house dust particles and diesel
exhaust particles as allergen carriers. Clin Exp Allergy, 28: 702-708.
Ostro BD, Broadwin R, Lipsett MJ. 2000. Coarse and fine particles and daily mortality in
the Coachella Valley, California: A follow-up study. J Expo Anal Environ Epidemiol 10:
412-419.
Ostro BD, Sanchez JM, Aranda C, Esekland GS. 1996. Air pollution and mortality: results
from a study of Santiago, Chile. J Expo Anal Environ Epidemiol 6:97-114.
Ostro B, Chestnut L, Vichit-Vadakan N, Laixuthai A. 1999b. The impact of particulate
matter on daily mortality in Bangkok, Thailand. J Air Waste Manag Assoc 49: 100-107.
Ostro BD, Lipsett MJ, Mann JK, Braxton-Owens H, White MC. 1995. Air pollution and
asthma exacerbations among African-American children in Los Angeles. Inhalation Toxicol
7: 711-722.
Ostro BD, Eskeland G, Sanchez JM, Feyzioglu T. 1999a. Air pollution and health effects:
A study of medical visits among children in Santiago, Chile. Environ Health Perspect 107:
69-73.
Otha K, Yamashita N, Tajima M, Miyaska T, Nakano J, Nakajima M, Ishii A, Horiuchi T,
Mano K, Miyamoto T. 1999. Diesel exhaust particulate induces airway hyperresponsiveness
in murine model: Essential role of GM-CSF. J Allergy Clin Immunol, 104: 1024-1030.
emission.
Özkaynak H, Xue J, Spengler J, Wallace L, Pellizzari E, Jenkens P. 1996. Personal exposure
to airborne particles and metals: Results from the particle team study in Riverside,
California. Journal of Exposure Analysis and Environmental Epidemiology, 6(1): 57-78.
Özkaynak H. et al. 1995. Association between daily mortality and air pollution in Toronto,
Canada. Proceedings of the International Society for Environmental Epidemiology.
Noordwijkerhourt, The Netherlands. August, 1995.
Pandya RJ, Solomon G, Kinner A, Balmes JR. 2002. Diesel exhaust and asthma: hypotheses
and molecular mechanisms. Environ Health Perspect 110 (suppl 1): 103-112.
Patterson E, Eatough DJ. 2000. Indoor/outdoor relationships for ambient PM2.5 and
associated pollutants: epidemiological implications in Lindon, Utah. J Air & Waste Manage
Assoc 50: 103-110.
Pekkanen J, Brunner EJ, Anderson HR, Tiittanen P, Atkinson RW. 2000. Daily
concentrations of air pollution and plasma fibrinogen in London. Occup Environ Med 57:
818-822.
Pekkanen J, Timonen K.L, Ruuskanen J, Reponen A, Mirme A. 1997. Effects of ultrafine
and fine particles in urban air on peak expiratory flow among children with asthmatic
symptoms. Environ Res 74:24-33.
Perera FP, Jedrychowski W, Rauh V, Whyatt RM. 1999. Molecular epidemiologic research
on the effects of environmental pollutants on the fetus. Environ Health Perspect, 107
(suppl3): 451-460.
Peters JM, Avol E, Navidi W, London SJ, Gauderman J, Lurmann F, Linn WS, Margolis H,
Rappaport E, Gong H, Thomas DC. 1999b. A study of twelve Southern California
Communities with differing levels and types of air pollution. I. Prevalence of respiratory
morbidity. Am J Respir Care Med, 159: 760-767.
Peters A, Frohlich M, Döring A, Immervoll T, Wichmann H-E, Hutchinson WL, Pepys MB,
Koenig W. 2001a. Particulate air pollution is associated with an acute phase response in
men, results from the MONICA-Augsburg Study. Eur Heart J 22: 1198-1204.
Peters A, Liu E, Verrier RL, Schwartz J, Gold DR, Mittleman M, Baliff J, Oh JA, Allen G,
Monahan K, Dockery DW. 2000a. Air pollution and incidence of cardiac arrhythmia.
Epidemiology 11: 11-17.
Peters A, Dockery DW, Muller JE, Mittleman MA. 2001b. Increased particulate air
pollution and the triggering of myocardial infarction. Circulation 103: 2810-2815.
emission.
Peters JM, Avol E, Gauderman J, Linn WS, Navidi W, London SJ, Margolis H, Rappaport
E, Vora H, Gong H, Thomas DC. 1999a. A study of twelve Southern California
Communities with differing levels and types of air pollution.II. Effects on pulmonary
function. Am J Respir Care Med, 159: 768-775.
Peters A, Döring A, Wichmann H-E, Koenig W. 1997c. Increased plasma viscosity during
an air pollution episode: A link to mortality? Lancet 349: 1582-1587.
Peters A, Dockery DW, Heinrich J, Wichmann HE. 1997b. Short-term effects of particulate
air pollution on respiratory morbidity in asthmatic children. Eur Respir J, 10: 872-879.
Peters A, Liu E, Verrier RL, Schwartz J, Gold DR, Mittleman M, Baliff J, Oh JA, Allen G,
Monahan K, Dockery DW. 2000a. Air pollution and incidence of cardiac arrhythmia.
Epidemiology 11: 11-17.
Peters A, Skorkovsky J, Kotesovec F, Brynda J, Spix C, Wichmann HE, Heinrich J. 2000b.
Association between mortality and air pollution in Central Europe. Environ Health Perspect,
108: 283-287
Peters A, Wichmann E, Tuch T, Heinrich J. and Heyder J. 1997a. Respiratory effects are
associated with the number of ultrafine particles. Am J Resp Crit Care Med 155:1376-1383.
Poloniecki JD, Atkinson RW, de Leon AP, Anderson HR. 1997. Daily time series for
cardiovascular hospital admissions and previous day's air pollution in London, UK. Occup
Environ Med 54: 535-540.
Pope III CA, Verrier RL, Lovett EG, Larson AC, Raizenne ME, Kanner RE, Schwartz J,
Villegas GM, Gold DR, Dockery DW. 1999b. Heart rate variability associated with
particulate air pollution. Am Heart J 138 (5 Pt 1): 890-899.
Pope III CA, Hill RW, Villegas GM. 1999a. Particulate air pollution and daily mortality on
Utah's Wasatch Front. Environ Health Perspect 107: 567-573.
Pope III CA. 2000. Epidemiology of fine particulate air pollution and human health:
biologic mechanisms and who’s at risk. Environ Health Perspect 108 Suppl 4: 713-723.
Pope III CA, Dockery D.W, Schwartz J. 1995a. Review of epidemiological evidence of
health effects of particulate air pollution. Inhalation Toxicol 7:1-18.
Pope III CA, Burnett RT, Thun MJ, Calle EE, Krewski D, Ito K, Thurston GD. 2002. Lung
cancer, cardiopulmonary mortality, and long-term exposure to fine particulate air pollution.
JAMA 287: 1132-1141.
emission.
Pope III CA, Verrier RL, Lovett EG, Larson AC, Raizenne ME, Kanner RE, Schwartz J,
Villegas GM, Gold DR, Dockery DW. 1999b. Heart rate variability associated with
particulate air pollution. Am Heart J 138: 890-899.
Pope III CA, Hill RW, Villegas GM. 1999a. Particulate air pollution and daily mortality on
Utah’s Wasatch Front. Environ Health Perspect, 107: 567-573.
Pope III CA, Thun MJ, Namboodiri MM, Dockery DW, Evans JS, Speizer FE, Heath CW.
1995b. Particulate air pollution as a predictor of mortality in a prospective study of U.S.
adults. Am J Respir Crit Care Med, 151: 669-674.
Pope III CA, Dockery D.W, Schwartz J. 1995a. Review of epidemiological evidence of
health effects of particulate air pollution. Inhalation Toxicol 7:1-18.
Pope III CA, Thun MJ, Namboodiri MM, Dockery DW, Evans JS, Speizer FE, Heath CW.
1995b. Particulate air pollution as a predictor of mortality in a prospective study of U.S.
adults. Am J Respir Crit Care Med, 151: 669-674.
Pope CA 3rd, Burnett RT, Thun MJ, Calle EE, Krewski D, Ito K, Thurston GD. 2002. Lung
cancer, cardiopulmonary mortality and long-term exposure to fine particulate air pollution.
JAMA 287: 1132-1141.
Pope III CA, Dockery D. W, Spengler J. D, Raizenne M. E. 1991. Respiratory health and
PM10 pollution: a daily time series analysis. Am. Rev. Respir. Dis. 144: 668-674.
Pope III CA. 1991. Respiratory hospital admissions associated with PM10 pollution in Utah,
Salt Lake and Cache Valleys. Arch Environ Health 46(2):90-97.
Pope III CA and Dockery DW. 1992. Acute health effects of PM10 pollution on symptomatic
children. Am. Rev. Respir. Dis. 145: 1123-1128.
Pope III CA, Kalkstein LS. 1996. Synoptic weather modeling and estimates of the
exposure-response relationship between daily mortality and particulate air pollution. Environ
Health Perspect 104:414-420.
Quénel P, Zmirou D, Dab W, LeTertre A, Medina S. 1999. Premature deaths and long-term
mortality effects of air pollution. (Letter. Int J Epidemiol 28: 362.
Renwick LC, Donaldson K, Clouter A. 2001. Impairment of alveolar macrophage
phagocytosis by ultrafine particles. Toxicol Appl Pharmacol 172: 119-127.
emission.
Reynolds P, von Behren J, Gunier RB, Goldberg DE, Hertz A, Smith DF. 2002. Childhood
cancer incidence rates and hazardous air pollutants in California: An exploratory analysis.
Environ Health Perspect in press.
Rittinghausen S, Mohr U, Dungworth DL. 1997. Pulmonary cystic keratinizing squamous
cell lesions of rats after inhalation/instillation of different particles. Exp Toxic Pathol, 49:
433-446.
Ritz B, Yu F, Chapa G, Fruin S. 2000. Effect of air pollution on preterm birth among
children born in Southern California between 1989 and 1993. Epidemiology 11: 502-511.
Rudell B, Ledin MC, Hammarstrom U, Stjernberg N, Lundback B, Sandstrom T. 1996.
Effects on symptoms and lung function in humans experimentally exposed to diesel exhaust.
Occup Environ Med, 53: 658-662.
Salvi S, Blomberg A, Rudell B, Kelly F, Sandstrom T Holgate ST, Frew A. 1999. Acute
inflammatory responses in the airways and peripheral blood after short-term exposure to
diesel exhaust in healthy human volunteers. Am J Respir Crit Care Med;159(3):702-9
Salvi S, Blomberg A, Rudell B, et al. 1997. Acute inflammatory responses in the airways
and peripheral blood following short term exposure to diesel exhaust in healthy human
volunteers. Am J Respir Crit Care Med, 155: 425 (Abstract).
Salvi S and Holgate ST. 1999. Mechanisms of particulate matter toxicity. Clinical and
Experimental Allergy, 29: 1187-1194.
Samet JM, Zeger SL, Dominici F, Curriero F, Coursac I, Dockery DW, Schwartz J,
Zanobetti A. 2000b. The National Morbidity, Mortality, and Air Pollution Study:Morbidity
and Mortality from Air Pollution in the United States. Health Effects Research Report
Number 94, Part 2, June. 2000. Health Effects Institute, Cambridge, MA. 93pp
Samet JM, Dominici F, Curriero FC, Coursac I, Zeger SL. 2000a. Fine particulate air
pollution and mortality in 20 U.S. cities, 1987-1994. N Engl J Med 343(24): 1742-1749.
Samet JM, Zeger SL, Dominici F, Curriero F, Coursac I, Dockery DW, Schwartz J,
Zanobetti A. 2000b. The National Morbidity, Mortality, and Air Pollution Study:Morbidity
and Mortality from Air Pollution in the United States. Health Effects Research Report
Number 94, Part 2, June (2000. Health Effects Institute, Cambridge, MA. 93pp
Samet JM, Zeger SL, Berhane K. 1995. The association of mortality and particulate air
pollution. In Particulate Air Pollution and Daily Mortality: Replication and Validation of
emission.
Selected Studies (The Phase I. Report of the Particle Epidemiology Evaluation Project)
Health Effects Institute, Cambridge, MA, pp. 3-104.
Samet JM, Zeger SC, Kelsall JE, Xu J, Kalkstein LS. 1997. Weather, air pollution and
mortality in Philadelphia. 1973-1980. In Particulate Air Pollution and Daily Mortality:
Analyses of the Effects of Weather and Multiple Air Pollutants (The Phase I.B Report of the
Particle Epidemiology Evaluation Project) Health Effects Institute, Cambridge, MA, pp.
1-30.
Samet JM, Zeger SL, Kelsall, Xu J, Kalksein. 1998. Does weather confound or modify the
association of particulate air pollution with mortality? Environ Res, 77: 9-19.
Samet JM, Dominici F, Curriero FC, Coursac I, Zeger SL. 2000a. Fine particulate air
pollution and mortality in 20 U.S. cities, 1987-1994. N Engl J Med 343(24): 1742-1749.
Schwartz J. 2000b. Harvesting and long-term exposure effects in the realtion between air
pollution and mortality. Am J Epidemiol 151: 440-448.
Schwartz J. 1992. Particulate air pollution and daily mortality in Steubenville, Ohio. Am J
Epidemiol 135:12-19.
Schwartz J, Dockery DW. 1992. Increased mortality in Philadelphia associated with daily
air pollution concentrations. Am Rev Respir Dis 145:600-604.
Schwartz J. 1993. Air pollution and daily mortality in Birmingham, Alabama. Am J
Epidemiol 137: 1136-1147.
Schwartz J. 1994a. Total suspended particulate matter and daily mortality in Cincinnati,
Ohio. Environ Health Perspect 102:186-189.
Schwartz J. 1994b. Air pollution and hospital admissions for the elderly in Birmingham,
Alabama. Am J Epidemiol 139(6):589-598.
Schwartz J. 1999. Air pollution and hospital admissions for heart disease in eight U.S.
counties. Epidemiology, 10: 17-22.
Schwartz J, Dockery DW, Neas LM, Wypij D, Ware JH, Spengler JD, Koutrakis P, Speizer
FE, Ferris BG. 1994. Acute effects of summer air pollution on respiratory symptom
reporting in children. Am J Respir Crit Care Med 150: 1234-1242.
Schwartz J, Dockery DW, Neas LM. 1996. Is daily mortality associated specifically with
fine particles? J. Air Waste Manag. Assoc. 46: 927-939.
emission.
Schwartz J. 1997. Air pollution and hospital admission for cardiovascular disease in Tucson.
Epidemiology, 8: 371-377.
Schwartz J. 1999. Air pollution and hospital admissions for heart disease in eight U.S.
counties. Epidemiology, 10: 17-22.
Schwartz J, Norris G, Larson T Sheppard L, Claiborne C, Koenig J. 1999. Episodes of high
coarse particle concentration are not associated with increased mortality. Environ Health
Perspect:107 (5): 339-342.
Schwartz J. 2000a. Assessing confounding, effect modification, and thresholds in the
association between ambient particles and daily deaths. Environ Health Perspect 108: 563568.
Schwartz J. 2000b. Harvesting and long term exposure effects in the realtion between air
pollution and mortality. Am J Epidemiol 151: 440-448.
Schwartz J. 2001. Air pollution and blood markers of cardiovascular risk. Environ Health
Perspect 109(Suppl 3): 405-409.
Schwartz J, Laden F, Zanobetti A. 2002. The concentration-response relation between PM2.5
and daily deaths. Environ Health Perspect 110: 1025-1029.
Schwartz J, Laden F, Zanobetti A. 2002. The Concentration-Response Relation between
PM2.5 and Daily Deaths. Environ Health Perspect 110: 1025-1029.
Schwartz J. 1994c. What are people dying of on high air pollution days? Environ Res 64:
26-35.
Seaton A, Soutar A, Crawford V, Elton R, McNerlan S, Cherrie J, Watt M, Agius R, Stout
R. 1999. Particulate air pollution and the blood. Thorax 54: 1027-1032.
Sheppard L, Levy D, Norris G, Larson TV, Koenig JQ. 1999. Effects of ambient air
pollution on nonelderly asthma hospital admissions in Seattle, Washington 1987-1994.
Epidemiology 10 (1): 23-30.
Simpson RW, Williams G, Petroeschevsky A, Morgan G, Rutherford S. 1997. Associations
between outdoor air pollution and daily mortality in Brisbane, Australia. Arch Environ
Health 52:442-454.
emission.
Sioutas C, Koutrakis P, Ferguson ST, Burton RM. 1995. Development and evaluation of a
prototype ambient particle concentrator for inhalation exposure studies. Inhal Toxicol 7:
633-644.
Steerenberg PA, Zonnenberg JA, Dormans JA, Joon PN, Wouters IM, van Bree L, Scheepers
PT, Van Loveren H. 1998. Diesel exhaust particles induced release of interleukin 6 and 8
by (primed) human bronchial epithelial cells (BEAS 2B) in vitro. Exp Lung Res 4: 85-100
Sunyer J, Spix C, Quenel P, Ponce-de-Leon A, Ponka A, Barumandzadeh T, Touloumi G,
Bacharova L, Wojtyniak B, Vonk J, Bisanti L, Schwartz J, Katsouyanni K. 1997. Urban air
pollution and emergency admissions for asthma in four European cities: the APHEA Project.
Thorax 52: 760-765.
Suwa T, Hogg JC, Quinlan KB, Ohgami A, Vincent R, van Eeden SF. 2002. Particulate air
pollution induces progression of coronary atherosclerosis. J Am Coll Cardiol 39: 935–942
Takano H, Ichinose T, Miyabara Y, Yoshikawa T, Sagai M. 1998. Diesel exhaust particles
enhance airway responsiveness following allergen exposure in mice. Immunopharmacol
Immunotoxicol 20: 329-336.
Takano H, Yoshikawa T, Ichinose T, Miyabara Y, Imaoka K, Sagai M. 1997. Diesel exhaust
particles enhance antigen-induced airway inflammation and local cytokine expression in
mice. Am J Respir Crit Care Med 156: 36-42.
Takenaka H, Zhang K, Diaz-Sanchez D, Tsien A, Saxon A. 1995. Enhanced human IgE
production results from exposure to the aromatic hydrocarbons from diesel exhaust: Direct
effects on B-cell IgE production. J Allergy Clin Immunol 95: 103-115.
Tancell PJ, Rhead MM, Trier CJ, et al. 1995. The sources of benzo[a]pyrene in diesel
exhaust emissions. Sci Total Environ 162:179-186. (Cited in U.S. EPA, 2002b)
Terashima T, Wiggs B, English D, Hogg JC, van Eeden SF. 1997. Phagocytosis of small
carbon particles (PM10) by alveolar macrophages stimulates the release of
polymorphonuclear leukocytes from bone marrow. Am J Respir Crit Care Med
155:1441-1447.
Tsien A, Diaz-Sanchez D, Ma J, Saxon A. 1997. The organic component of diesel exhaust
particles and phenathrene, a major polyaromatic hydrocarbon constituent, enhances IgE
production by IgE-secreting EBV-transformed human B cells in vitro. Toxicol Appl
Pharmacol 142: 256-263.
emission.
U.S. EPA. 1998a. SAB Review Draft Health Assessment Document of Diesel Emissions.
EPA/600/8-90/057C. United States Environmental Protection Agency, Office of Research
and Devlopment, Washington DC 20460.
U.S. EPA. 1996. Air Quality Criteria for Particulate Matter, Volumes 1-3. U.S.
Environmental Protection Agency (U.S. EPA), Office of Research and Development,
National Center for Environmental Assessment, Research Triangle Park, NC.
EPA/600/P-95/001a-cf.
U.S. EPA. 2002a. Third External Review Draft of Air Quality Criteria for Particulate
Matter: Volumes I & II. National Center for Environmental Assessment-RTP Office, Office
of Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC. EPA/600/P-99/002aC, April 2002.
U.S. EPA . 2002b. Health assessment document for diesel engine exhaust. Prepared by the
National Center for Environmental Assessment, Washington, DC, for the Office of
Transportation and Air Quality, EPA/600/8-90/057F. Available from: National Technical
Information Service, Springfield, VA, PB2002-107661, and <http://www.epa.gov/ncea>.
U.S. EPA. 1994. Methods for derivation of inhalation reference concentrations and
application of inhalation dosimetry. EPA/600/8-90/066F. Research Triangle Park, NC,
Office of Research and Development.
U.S. EPA. 1987. June-September, 6-9 AM, ambient air benzene concentrations in 39 U.S.
cities, 1984-1986. Research Triangle Park, NC: U.S. Environmental Protection Agency,
Atmospheric Sciences Research Lab. EPA/600/D-87/160.
U.S. EPA. 1999. U.S. Environmental Protection Agency 1999 National Air Toxics Program:
The Integrated Urban Strategy, Notice. Federal Register, 64 (137): 38706-38740
U.S. EPA. 1998b. Emission Factor Documentation For AP-42. Section 10.6.1, prepared for
the U. S. Environmental Protection Agency, OAQPS/EFIG, by Midwest Research Institute,
Cary, NC, December 1998.
U.S. EPA. 2002. Health Assessment Document for Diesel Engine Exhaust. EPA/600/890/057F. National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Washington, DC.
Valberg PA, Watson AY. 1998. Alternative hypotheses linking outdoor particulate matter with
daily morbidity and mortality. Inhalation Toxicology 10: 641-662.
emission.
Valberg PA, Crouch EAC. 1999. Meta-analysis of rat lung tumors from lifetime inhalation of
diesel exhaust. Environ Health Perspect 107: 693-699.
van den Eden S, Tager I, Quesenberry C, Mann J, Shan J, Lurnann F. and Lugg M. 1998.
Exposure to particulates and cardiovascular (CV) hospitalization among members of a large
health maintenance organization in California. Am J Resp Crit Care Med 157(3):A260.
van Vliet P, Knape M, de Hartog J, Jansses N, Harssema H, Brunekreef B. 1997. Motor vehicle
exhaust and chronic respiratory symptoms in children living near freeways. Environ Res 74:
122-132.
Vedal S, Brauer M, White R, Petkau J. 2003. Air pollution and daily mortality in a city with low
levels of pollution. Environ Health Perspect. 111: 45-51.
Veronesi B, Oortgiesen M, Carter JD, Devlin RB. 1999. Particulate matter initiates
inflammatory cytokine release by activation of capsaicin and acid receptors in a human bronchial
epithelial cell line. Toxicology and Applied Pharmacology 154: 106-115.
Vincent R, Kumarathasan P, Goegan P, Bjarnason SG, Guénette J, Bérubé D, Adamson IY,
Desjardins S, Burnett RT, Miller FJ, Battistini B. 2001. Inhalation Toxicology of Urban
Ambient Particulate Matter: Acute Cardiovascular Effects in Rats. Research Report 104. Health
Effects Institute, Boston MA.
Watkinson WP, Campen MJ, Costa DL. 1998. Cardiac arrhythmia induction after exposure to
residual oil fly ash particles in a rodent model of pulmonary hypertension. Toxicol Sci 41: 209216.
Watson AY and Green GM. 1995. Noncancer effects of diesel emissions: animal studies. In:
Diesel Exhaust: A Critical Analysis of Emissions, Exposure and Health Effects. A Special
Report on the Institute’s Diesel Working Group. Health Effects Institute Cambridge, MA. 141164.
Wichmann H-E, Spix C, Tuch T, Wölke G, Peters A, Heinrich J, Kreyling WG, Heyder J. 2000.
Daily Mortality and Fine and Ultrafine Particles in Erfurt, Germany. Part I: Role of Particle
Number and Particle Mass. Research Report 98. Health Effects Institute, Cambridge MA.
Wichmann H-E, Spix C, Tuch T, Wölke G, Peters A, Heinrich J, Kreyling WG, Heyder J. 2000.
Daily Mortality and Fine and Ultrafine Particles in Erfurt, Germany. Part I: Role of Particle Number
and Particle Mass. Research Report 98. Health Effects Institute, Cambridge MA.
Williams PT, Andrews GE, Bartle KD. 1987. The role of lubricating oil in diesel particulate and
particulate PAH emissions. SAE Technical Paper Ser. No. 872084. (Cited in U.S. EPA, 2002b)
emission.
Williams DJ, Milne JW, Quigley SM, et al. 1989. Particulate emissions from in-use motor
vehicles. II. Diesel vehicles. Atmos Environ 23(12):2647. (Cited in U.S. EPA, 2002b)
Wilson WE, Suh HH. 1997. Fine particles and coarse particles: concentration relationships
relevant to epidemiologic studies. J Air Waste Manage Assoc 47: 1238-1249.
Woodruff TJ, Grillo J, Schoendorf KC. 1997. The relationship between selected causes of
postneonatal infant mortality and particulate air pollution in the United States. Environ Health
Perspect 105: 608-612.
Yang W, Jennison BL, Omaye ST. 1997. Air pollution and asthma emergency room visits in
Reno Nevada. Inhal Toxicol 9: 15-29.
Zeger SL, Thomas D, Dominici F, Samet JM, Schwartz J, Dockery D, Cohen A. 2000.
Exposure measurement error in time-series studies of air pollution: Concepts and consequences.
Environ Health Perspect 108: 419-426.
Zeger SL, Thomas D, Dominici F, Samet JM, Schwartz J, Dockery D, Cohen A. 2000.
Exposure measurement error in time-series studies of air pollution: Concepts and consequences.
Environ Health Perspect 108: 419-426.
Zeger S, Dominici F, Samet J. 1999. Harvesting-resistant estimates of air pollution effects on
mortality. Epidemiology 10: 171-175.
Zemp E, Elsasser S, Schindler C, Künzli N, Perruchoud AP, Domenighetti G, Medici T,
Ackermann-Liebrich U, Leuenberger P, Monn C, Bolognini G, Bongard JP, Brändli O, Karrer
W, Keller R, Schöni MH, Tschopp JM, Villiger B, Zellweger JP, and the SAPALDIA Team.
1999. Long-term ambient air pollution and respiratory symptoms in adults (SAPALDIA Study)
Am. J. Respir. Crit. Care Med. 159: 1257-1266.
emission.

Table of Contents

Transcription

Similar documents

Application Sheets

Flame Photometer (Model ESICO- 1381) for determination of alkali

PARTICULATE MATTER AND CORONARY ARTERY DISEASE

Testimony of Tee Guidotti, MD, MPH, DABT Representing the American Thoracic Society

How Air Pollution Affects Your Health Air Pollution Affects Your

Press release – 04/10/05 IONISEUR-F4: equipment for

Position Paper on the proposed EU directive on Particulate Matter Introduction

original PDF - European Geosciences Union

Air pollution Preamble

campbellton, nb - Newspaper Toolbox