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Essay On Troposphere Pollution Shanghai

Abstract

Ambient daytime and nighttime PM2.5 (particulate matter with aerodynamic diameter less than 2.5 μm) and TSP (the total suspended particulates) samples were collected at two sites (named Pudong and Jinshan) in Shanghai. The concentrations of PM2.5 and TSP were lower at Pudong than at Jinshan. Higher PM2.5 and TSP concentrations were observed during daytime than nighttime for both sites. Carbonaceous aerosol and secondary sulfate were the most abundant components. Larger enrichment factor (EFs) of Zn, Pb, Cl, and S for Jinshan nighttime were observed than for other sampling periods. PM2.5 showed higher relative spatial uniformity (the coefficients of divergence, COD = 0.18) than TSP (COD = 0.23) during the sampling period. The variations of chemical components and the species ratios showed that the contributions of primary particulate emissions in Jinshan (industrial zone) were more significant than in Pudong (residential zone). View Full-Text

Keywords: PM2.5; TSP (the total suspended particulates); carbonaceous fractions; ions; elements; ShanghaiPM2.5; TSP (the total suspended particulates); carbonaceous fractions; ions; elements; Shanghai

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DOI: 10.1039/C4PP90037E (Perspective) Photochem. Photobiol. Sci., 2015, 14, 149-169

Changes in air quality and tropospheric composition due to depletion of stratospheric ozone and interactions with changing climate: implications for human and environmental health

Received 20th October 2014 , Accepted 20th October 2014

First published on 7th November 2014


UV radiation is an essential driver for the formation of photochemical smog, which includes ground-level ozone and particulate matter (PM). Recent analyses support earlier work showing that poor outdoor air quality is a major environmental hazard as well as quantifying health effects on regional and global scales more accurately. Greater exposure to these pollutants has been linked to increased risks of cardiovascular and respiratory diseases in humans and is associated globally with several million premature deaths per year. Ozone also has adverse effects on yields of crops, leading to loss of billions of US dollars each year. These detrimental effects also may alter biological diversity and affect the function of natural ecosystems. Future air quality will depend mostly on changes in emission of pollutants and their precursors, but changes in UV radiation and climate will contribute as well. Significant reductions in emissions, mainly from the energy and transportation sectors, have already led to improved air quality in many locations. Air quality will continue to improve in those cities/states that can afford controls, and worsen where the regulatory infrastructure is not available. Future changes in UV radiation and climate will alter the rates of formation of ground-level ozone and photochemically-generated particulate matter and must be considered in predictions of air quality. The decrease in UV radiation associated with recovery of stratospheric ozone will, according to recent global atmospheric model simulations, lead to increases in ground-level ozone at most locations. If correct, this will add significantly to future ground-level ozone trends. However, the spatial resolution of these global models is insufficient to inform policy at this time, especially for urban areas. UV radiation affects the atmospheric concentration of hydroxyl radicals, ˙OH, which are responsible for the self-cleaning of the atmosphere. Recent measurements confirm that, on a local scale, ˙OH radicals respond rapidly to changes in UV radiation. However, on large (global) scales, models differ in their predictions by nearly a factor of two, with consequent uncertainties for estimating the atmospheric lifetime and concentrations of key greenhouse gases and air pollutants. Projections of future climate need to consider these uncertainties. No new negative environmental effects of substitutes for ozone depleting substances or their breakdown-products have been identified. However, some substitutes for the ozone depleting substances will continue to contribute to global climate change if concentrations rise above current levels.


Introduction

The degradation of air quality is one of the major environmental hazards facing modern society. Human activities result in the emission of many chemicals to the atmosphere, which are either toxic themselves, or produce noxious compounds when exposed to ambient ultraviolet (UV) radiation. UV radiation is an essential driver for the generation of ground-level ozone (O3) and some particulate matter (PM, frequently called aerosol) including sulfate, nitrate, and organic aerosols. These pollutants have major health implications for humans and the environment. Future changes in tropospheric UV radiation, whether from stratospheric ozone changes or other factors such as clouds, are likely to contribute to trends in air quality and associated health effects.

UV radiation makes hydroxyl (˙OH) radicals, the so-called cleaning agents of the troposphere. These radicals limit the atmospheric lifetime of many gases that are important to both tropospheric and stratospheric chemistry as well as climate change, including methane (CH4), hydrogen-containing halocarbons (e.g. hydrofluorocarbons, hydrochlorofluorocarbons and hydrobromocarbons), and the oxides of sulfur and nitrogen (SO2 and NOx).

This paper provides an assessment of how UV radiation affects air quality, particularly ground-level O3 and PM, in the context of the growing body of knowledge on their health impacts, geographic distributions, and long-term trends. It should be recognised that many factors drive these trends, including changes in both natural and anthropogenic emissions, as well as climate variability through changes in temperature, moisture, and atmospheric circulation patterns. Tropospheric UV radiation is one of these factors, and its effects can be approximately superimposed onto the effects of the other factors, but complex non-linear interactions must be considered to obtain reliable estimates.

Since the previous assessment of the interactions between ozone depletion, climate change, and air-quality in the troposphere,10,11 significant advances are noted in: (i) understanding and quantifying the important consequences of poor air quality for human health, separating the effects of O3 and PM; (ii) understanding changes in air pollution on urban and regional scales, in terms of changes in anthropogenic emissions (increases or decreases, depending on location) as well as long-range transport; (iii) understanding long-term changes of key tropospheric oxidants (O3 and ˙OH) on continental and global scales, as anthropogenic emissions continue in an environment where both stratospheric ozone and climate are also changing.

An equally important advance is a better understanding of uncertainties inherent in numerical models used to predict the future chemical composition of the atmosphere. These computer models endeavor to represent and integrate the many chemical, physical, and biological processes that control air quality as well as climate. Recent inter-comparisons among the models (see below) highlight important differences that cast some doubt on the reliability of future projections, while also indicating a path to model improvements.

Ozone-depleting substances (ODS) could also affect air quality. While the long-lived chlorofluorocarbons (CFCs) break down almost exclusively in the stratosphere, the halogenated replacements break down in the troposphere. The cycling of the halogenated species in the troposphere needs to be assessed to ensure that there are no other significant short- and long-term effects that will result from the replacements for the halocarbons. Health effects could result from exposure to these substances, and are therefore included in this assessment (previously this was included in the health assessment, which now focuses on UV-mediated effects).

This paper provides summaries of the state of knowledge on ground-level O3, PM, and ˙OH radicals, and addresses some of the more complex but still largely unquantified interactions between air quality, climate change, and human activities. An important additional interaction between air quality and stratospheric ozone depletion is the introduction of substitute compounds for ODSs pursuant to international agreements. The last part of this paper provides an update on selected substitutes whose potential environmental and health impacts should be considered.

In summary, our assessment updates and reinforces several key conclusions. Air pollution is increasingly recognised as a major environmental hazard and a risk to human health, globally leading to several million premature deaths per year. Air pollution also damages vegetation and reduces agricultural yields, with associated economic losses estimated as $10–20 billion annually. UV radiation is an essential ingredient for the formation of ground-level O3 and some PM, and of ˙OH radicals that control the global self-cleaning capacity of the troposphere. Future trends in UV radiation will modulate future trends in air quality. Air quality is sensitive to other changes in the environment including atmospheric circulation, hydrological cycles, and temperatures, all of which are likely to change due to the combined effects of changing stratospheric ozone and climate. No new negative environmental effects of the substitutes for the ODSs have been identified.

Ground-level ozone

Health effects. Tropospheric O3 has significant effects on human morbidity and mortality. Premature mortality has been estimated in recent studies, which are summarised in Table 1. Ozone and particulate matter (PM) often co-occur in the troposphere and therefore their effects on human health are difficult to separate. However, there appears to be no interaction between these in terms of premature mortality. Earlier epidemiological studies (reviewed in ref. 12) have supported this conclusion and further studies by the same authors13,14 on the individual components of PM have shown that there is no interaction between these and ozone. Thus for the purposes of protecting human health, PM and O3 can be treated separately. Premature mortality associated with exposure to ground-level ozone, while lower than that from PM, is still substantial with several hundred thousand people affected globally each year (Table 1). Recent studies are broadly consistent on effects of ozone on mortality in humans from cardiovascular and respiratory diseases in Oporto, Portugal,15 Taipei,16 and Prague.17,18

Table 1Premature mortality from ground-level ozone (O3) and particulate matter (PM)



Ozone-related morbidities manifested as acute and chronic bronchitis, asthma and/or atopic dermatitis,22 appendicitis,23 venous thromboembolic disease, and pulmonary embolisms24 have been reported. Ozone may also interact synergistically with viral infections. A study in Hong Kong reported interactive effects between viral infections and concentration of O3 that resulted in increased risks for hospitalisation for respiratory disease.25 The mechanism for this apparent synergism was not reported.

A number of studies have extrapolated the effects of exposure to O3 into the future. Based on the OECD19 study, premature deaths from ground-level ozone will increase to about 0.75 million per year worldwide by 2050. The greatest increases are predicted in India (130 premature deaths per million per decade by 2050) but those in OECD countries will be almost as large (95 premature deaths per million per decade by 2050), mostly as a result of greater sensitivity in an ageing population.19 Modelling of the interactions between concentrations of tropospheric ozone, precursors of ozone in the atmosphere, and climate change suggests that, by 2050, concentrations of O3 will increase in developing countries and decrease in developed countries,26 due to regional differences in future emissions.

SourceYearAreaPremature mortality (millions per year)
From O3From PM
OECD192010Global0.351.4
2050Global0.753.6
Lim et al.182010Global0.05–0.272.8–3.6
Fang et al.202000Global0.381.5
Fann et al.212005US only0.0050.05–0.2
Effects on plants. For plants, the most important air-pollutant is O3;27,28 particulates have not been observed to have substantial direct effects on plants. In our previous assessment,11 we noted that damage to crops by air-pollutants is likely to become more severe in the future. Since the last assessment, further studies have reinforced this conclusion. Based on a scenario of a world population of 9.1 billion, concentration of CO2 of 550 ppm,† a concentration of ozone of 60 ppb (about 10 ppb above current), and the climate warmer by ca. 2 °C by 2050, Jaggard et al.29 postulated that yields of major crops (e.g. wheat, rice, soy, and maize) would be reduced by about 5% because of O3. However, this may be compensated by an increase in yield for most crops by about 13% because of the increased concentrations of CO2, depending on water availability.

In a review of studies on the effects of O3 on plants, it was concluded that reductions in the yields of 18 to 27% result from exposure to O3 at concentrations of 70 to 100 ppb, at the upper end of typical regional concentrations.30 Not all crops are equally sensitive to ozone. In a sensitive crop, such as soybean, yields in several cultivars were shown to decrease by as much as a factor of 2 with long-term exposure to 20–30 ppb (24 hour mean) of added ozone.31 Reductions in photosynthesis upon exposure to O3 were estimated in three different forest types and found to be statistically significant for an orange orchard, and observable for ponderosa pine.32 The potential for damage from O3 in crop plants was assessed in relation to the IPCC (pessimistic) A2 scenario33 for 2100.1 Changes in gross plant productivity resulting from changes in tropospheric O3 were projected to range from −40 to +15%, depending on location (Fig. 1). In another example, the ranges of global crop losses for wheat and soybean in 2030 as estimated from the IPCC A2 scenario were 5.4–26% and 15–19%, respectively.34 In this same study, yield reductions in the B1 (optimistic) scenario were 4.0–17% for wheat and 9.5–15% for soybean, with monetised annual losses estimated to range from $12–21 billion (year 2000 dollar equivalents).


Fig. 1 Global assessment of the projected percentage changes in gross primary productivity (GPP) due to O3 under the Intergovernmental Panel on Climate Change A2 scenario in 2100 within the World Wildlife Foundation Global 200 priority conservation areas. From Ainsworth et al.1 Reproduced with permission of the Royal Society.

Although most of the research on the effects of O3 in ecosystems has been directed toward plants, effects on soil microorganisms have also been reported.35 The abundance and diversity of methanogenic bacteria in soils of rice paddies were found to be reduced after exposure to elevated concentrations of O3 (60 ppb) in ground-level air.

Overall, productivity and yields of crops will likely be reduced as a result of increases in concentrations of tropospheric O3. There is some hope that genetic selection of plants tolerant to O3 will mitigate these adverse effects on production of food and fibre but other plants in the ecosystem are likely to suffer greater adverse effects. These are expected to impact diversity and functions of natural ecosystems. Other effects of climate change, e.g., mediated by temperature and precipitation, will also affect yields of crops.

Photochemical processes. Atmospheric ozone (O3) is generated primarily in the atmosphere by photochemical reactions involving UV radiation. In the stratosphere, it is made directly by the photo-dissociation of molecular oxygen (O2) into two oxygen atoms (2O), followed by the association of each of these atoms with remaining O2 (see Table 2) to make two O3 molecules. In the troposphere, this direct formation is not possible because photons of sufficient energy to dissociate O2 (λ < 240 nm) are nearly completely absorbed by stratospheric O2, and so are not available in the troposphere. Descent of stratospheric ozone to the troposphere does occur (e.g., during stratospheric intrusions common during springtime at mid-latitudes) but accounts for only about 1/10th of the tropospheric production in global models,6 and is only a minor source of ozone found near the surface in polluted regions.

Table 2Atmospheric photochemical reactions





Formation of O3 in polluted urban atmospheres has been recognised since the 1950s,36 and occurs when mixtures of volatile organic compounds (VOCs) and nitrogen oxides (NOx = NO + NO2) are exposed to the UV radiation available in the troposphere. The chemistry leading to tropospheric O3 formation is generally complex because it (i) requires the absorption of multiple photons, (ii) is augmented by several catalytic cycles, and (iii) can be fuelled by many different VOCs, including those of anthropogenic and biogenic origin. Table 2 provides a highly simplified schematic of the major chemical pathways. The production of tropospheric O3 is autocatalytic because photolysis of an initial amount of O3 (Table 2, reactions 3 and 4) results in two ˙OH radicals, each of which can then continue through the reaction sequence to regenerate more O3. Note that two photons are required for this process (Table 2, reactions 3 and 7) with their combined minimum energy being more than sufficient to break O2 directly.

Reaction 3 in Table 2, the photolysis of ozone to yield excited oxygen atoms O*, is the primary source of radicals (˙OH, HO2, and RO2) involved in production of O3. It is also very sensitive to the overhead ozone column (see Table 1 of McKenzie et al.37). Other sources of radicals include the photolysis of formaldehyde (CH2O), hydrogen peroxide (H2O2), and nitrous acid (HONO). These latter compounds are typically the products of previous chemical reactions initiated by O3 and ˙OH and so they are sensitive to the production of primary radicals (Table 2, reaction 3) and therefore to UV-B radiation as affected by changes in stratospheric ozone.

Stratospheric ozone formation:Reaction
O2 + hν (λ < 240 nm) → O + O(1)
O + O2 → O3(2)
Tropospheric ozone and ˙OH formation:
O3 + hν (λ < 330 nm) → O* + O2(3)
O* + H2O → ˙OH + ˙OH(4)
˙OH + VOC + O2 → HO2 (or organic analog RO2) + other products(5)
HO2 + NO → NO2 + ˙OH(6)
NO2 + hν (λ < 420 nm) → NO + O(7)
O + O2 → O3(2)
Secondary tropospheric radical sources:
CH2O + hν (λ < 340 nm) + 2 O2 → HO2 + HO2 + CO(8)
H2O2 + hν (λ < 350 nm) → ˙OH + ˙OH(9)
HONO + hν (λ < 395 nm) → ˙OH + NO(10)
Distributions and trends. Tropospheric ozone has increased since preindustrial times, mostly because of increasing anthropogenic emissions of the precursor gases, VOCs and NOx. Relative to that, the effects of changing the UV radiation environment, e.g., from mid-latitude stratospheric ozone depletion, have been smaller but remain important due to the very large number of people living in areas with poor air quality.

Urban ozone trends differ in different cities. Considerable progress has been made in reducing urban O3 in Europe, the United States, and some other locations.2,38–40Fig. 2 shows the reductions in ground level ozone achieved in Los Angeles and Mexico City over the past several decades. In Beijing, from 2005 to 2011, ground-based measurements give daytime average O3 increasing at 2.6 ppb (5%) per year.41 In comparison, increases of 3% per year over 2002–2010 have been reported for the tropospheric O3 column above Beijing.42 Other Asian cities showing increases include Hong Kong (0.55 ppb per year over 1994–2007), Seoul (about 5 ppb over 1991–2007), and Tokyo with a doubling of days with O3 exceedances (incidences where air quality standards are exceeded).40


Fig. 2 Improvements in air quality in Los Angeles and Mexico City. Plotted is the 3-year average of the 4th highest maximum ozone 8-hour average. (From Parrish et al.2).

Regional ozone (ozone averaged over large areas extending well beyond cities) is increasing at some locations, particularly in densely populated areas, for example 6–7% per decade in the Indo-Gangetic Plains.43 Regional production of ozone was also shown to be a factor limiting air quality improvements from local emission reductions during the Beijing Olympics.44 Background ozone continues to increase at many locations, such as central and northwestern Europe, but is decreasing in eastern and southwestern Europe.38,39 Increases of 0.25 ppb per year have been reported for ground-level ozone at Mace Head, Ireland.45 In the western U.S., mid-tropospheric (3–10 km) ozone has increased by 0.6 ppb per year in the springtime over 1995–2008, and may be indicative of long range transport.46 Changes in the seasonal cycle observed at various mid-latitude locations in the Northern Hemisphere are consistent with increasing emissions of precursors.47 On the other hand, a review of measurements made with balloon-borne instruments (ozone sondes) and surface observations at remote locations showed that most of the increases in both hemispheres occurred in the early part of the 20–40 year record, with more recent changes characterised by little or no increase.48

Changes in atmospheric circulation are also likely to have affected trends of tropospheric ozone. Some increases may have been due to changes in the rate of stratosphere–troposphere air exchange.49 Trends at Mauna Loa, Hawaii, were influenced by changes in circulation due to El Niño, transporting air masses from different regions of Asia to Hawaii.50

Prediction of tropospheric concentrations of O3via numerical models remains problematic, with large differences at regional38 and global scales.6,51 This is illustrated in Fig. 3, where several global models, represented by different lines, are seen to differ by as much as 25% for current conditions, and future predictions show a wider range depending on the choice of future scenario as well as model. However, it should be noted that future photochemical ozone formation will in any case depend to a large extent on details of future emissions, particularly those associated with different fuel choices (e.g. diesel, gasoline, or biofuels).


Fig. 3 Global tropospheric ozone burden simulated by different models. Drawn from data in Table 1 of Young et al.6 Thin lines are for individual model results, thick lines are multi-model averages, for two scenarios of future emissions, RCP2.6 (green) and RCP8.5 (red), as defined by the Intergovernmental Panel on Climate Change.9

The specific response of tropospheric O3 to future changes in stratospheric O3 was modeled by Zeng et al.52 and Zhang et al.53 with both studies showing large-scale increases in tropospheric O3, as a net result of slowing both production and loss in response to declining UV levels. However the low resolution of their models (several degrees latitude × longitude) is insufficient to discern urban effects where higher levels of NOx are expected to maintain an opposite (positive) relationship between UV radiation and ground-level ozone54,55 that would indicate improvement in air quality in response to recovery of stratospheric ozone. Global models also do not agree well with measurements of the background atmosphere (e.g.ref. 56), indicating that there is still significant work to be done in understanding this chemistry. These uncertainties make it difficult to identify precisely which geographic regions will experience decreases in tropospheric ozone as stratospheric ozone recovers, and which ones will suffer increases. Nevertheless, all models agree that over large regions tropospheric ozone will increase.

Nitrogen dioxide. The near-term outlook for tropospheric oxidants can also be surmised from satellite measurements of NO2 (see Fig. 4), a precursor of tropospheric ozone and an important pollutant in its own right. The geographic distribution agrees with the general understanding of major emission sources, particularly over the U.S., Europe, and East Asia. Trends, also derived from satellite-based observations, are shown in the lower panel for specific regions. Notably, decreases in NO2 are seen to have occurred over the U.S. and Europe, in accordance with NOx emission reduction policies, and consistent with the reductions in urban ozone reported for these regions. However, positive trends are noted for east-central China, the Middle East, and north-central India. It seems likely that such recent trends will also continue into the near-term future, with the associated expectation that ground-level ozone (and other photochemical pollutants) may increase in some areas and decrease in others.
Fig. 4 Tropospheric vertical column of nitrogen dioxide, a major air pollutant and precursor of tropospheric ozone. Top panel: global distribution averaged from May 2004 to April 2005.3 Bottom panel: region-specific trends.8

Particulate matter

Particulate matter (PM) in the atmosphere consists of small solid or liquid particles suspended in air, also called aerosols. The size of PM is recognised as important for health effects, with PM smaller than 2.5 μm (termed PM2.5) being inhaled deeper into lungs than larger particles, typically measured as all particles below 10 μm (PM10).
Health effects. Particulate matter in the troposphere causes significant adverse health effects. A large body of literature spanning decades of research has been reviewed and assessed confirming this causal relationship (e.g.ref. 57). Table 1 shows recent estimates of premature deaths from particulate matter. Recent studies have reported PM health effects from different locations. In the time-series analysis of hospitalisations for venous thromboembolic disease in Chile between 2000 and 2007 discussed for ozone above, an increase in concentration of PM2.5 of 20 μg m−3 resulted in a relative risk (RR) of 1.05 (95% CI 1.03–1.06) for all hospitalisations.24 A similar RR was observed for pulmonary embolism. In a study of 90 cities in China over 1981–2000, those north of the river Huai had total suspended particulates (PM of all sizes) higher by 100 μg m−3 (95% CI = 61–307) than those south of the river, and were associated with a decrease in life-expectancy of 3 years (95% CI 0.4–5.6).58 Much of the PM2.5 is due to particles generated in the atmosphere by UV-dependent photochemistry (e.g. sulfate, nitrate, and organics, see below), although other PM sources can be extremely important, e.g., biomass burning plumes reaching densely populated urban areas (e.g., van Donkelaar et al.59).

Where control of emissions into the troposphere has resulted in decreases in PM2.5, fewer health effects have been observed. In a study of life expectancy in 545 U.S. counties, reductions in PM2.5 of 10 μg m−3 from 2000 to 2007 were associated with an increase in mean life expectancy of 0.35 years (SD = 0.16 years).60

Overall, the global relevance of particulates to human health is very large, and substantial changes are expected to occur in response to changes in climate.61 Future predictions are uncertain due to limitations of atmospheric models and their assumptions62 and, specifically for human health effects, the difficulty to clearly separate effects of O3 and PM2.5.63 Future changes in aerosols are uncertain but may be substantial regionally. A multi-model analysis of past and future trends in aerosol, described in Fig. 8 of Bais et al.,64 indicates large changes in industrialised regions, particularly in China.

Effects on plants. Direct effects of PM on plants appear to be minor, for example through direct deposition of PM on foliage.57 However, two important indirect effects should be recognised. The first is an increase in diffuse visible radiation from the scattering of solar photons by aerosol particles, altering photosynthetic efficiency within partly shaded canopies. The second is the surface deposition of some aerosol chemicals, for example, the heavy metals Cu, Ni, and Zn, with potential effects on soil chemistry, microbial communities, and nutrient cycling.57

Atmospheric processes. Particles in the atmosphere include those emitted directly, such as wind-blown dust and soil, combustion-generated soot (black carbon), and salt from sea-spray, as well as those formed in situ by condensation of vapours, such as sulfates, nitrates, and many organics. The latter, secondary, aerosols depend on UV-initiated reactions of ˙OH radicals (see section on Global ˙OH models), and thus are likely to be affected by changes in stratospheric ozone. However, we note the absence of specific studies addressing how changes in UV associated with stratospheric O3 would affect the formation and removal of tropospheric particles.

The formation of sulfate and nitrate aerosols is well understood in terms of the ˙OH oxidation of SO2 and NO2 giving sulfuric and nitric acids, respectively. While the majority of this production occurs in the gas phase, the sulfate and nitrate condense rapidly to form particles, particularly if ammonia is present. Chemical reactions in cloud and rain water can also contribute.65

Considerable progress has been made recently in understanding secondary organic aerosols (SOA), which previous observations had shown to be seriously underestimated by models. While many details remain poorly understood, numerous studies support the basic conceptual model that hydrocarbons are oxidised (by ˙OH and NO3 radicals, and O3) into a myriad of heavier, more functionalised molecules as well as smaller fragments.66 Molecules with multiple functional groups (e.g., alcohols, ketones, aldehydes, organic acids, nitrates and peroxides) typically have lower vapour pressures and therefore are likely to condense onto particles. However, quantification remains a problem due to the large number of chemical species contributing to particle mass. Significant advances in modelling have been made by classifying these multifunctional compounds according to relevant properties, such as vapour pressure,67,68 solubility,69 oxidation state,70 atomic ratios (O, C, H, etc.),71,72 and carbon number and polarity.73 In practice, for ambient aerosols many of these properties are not known and therefore cannot be used to constrain predictions. However, these modelling frameworks now allow exploratory sensitivity analyses to help identify the most important processes for more accurate parameterisation.

Removal of aerosols from the atmosphere is poorly understood. Ultimately removal from the atmosphere occurs by wet or dry deposition. Incorporation of aerosol particles into raindrops (wet deposition) leads to lifetimes estimated to range from 0.5 to 2 weeks.74–76 Dry deposition of particles is generally slower.65,77

Distributions and trends. The global distribution of aerosols is shown in Fig. 5. Satellite observations and models agree on broad features, including the dust belt extending from N. Africa to S. Asia, biomass burning evident over tropical S. America, and high values over E. Asia. These optical depth values represent the entire aerosol vertical column and not necessarily those at ground level. Surface network data are available in many countries and have been used to show detailed geographical and seasonal distributions of major chemical constituents of collected particles (e.g.ref. 78 for the U.S.).
Fig. 5 Annual average aerosol optical depth at 550 nm from the MODIS and MISR satellite instruments (top and bottom, for the years 2004–2006) and models (middle, for the year 2000). From Schindell et al.4

Heavily populated urban locations are of special interest, and some recently reported measurements in megacities are summarised in Table 3. World Health Organization (WHO) guidelines are frequently exceeded by all cities listed. Reductions in PM concentrations are occurring in many cities, in some cases well-documented by long-term urban monitoring networks, and evidently related to emissions-lowering technologies for both fixed and mobile sources. However, measurements at many polluted locations are still sparse, making an assessment of trend difficult.

Table 3Concentrations of particulate matter in megacitiesa



Future concentrations of aerosols are subject to similar scenario assumptions as other pollutants.4,79,80 Globally averaged sulfate concentrations have already decreased over the last two decades and are expected to continue decreasing. Organic and black carbon are expected to continue to increase over the next few decades globally, but then decrease, with timing and magnitude depending on the specific scenario.

On smaller geographic scales, aerosol concentrations are sensitive to local and regional emissions, and may improve or worsen depending on regulatory strategies. For example, a multi-model analysis of past and future trends in aerosols, described in Fig. 8 of Bais et al.64 shows strong regional reductions of aerosol concentrations by 2090, particularly in China.

CityPM10 (μg m−3)PM2.5 (μg m−3)Measurement period
WHO guidelines2010Annual mean
502524-hour mean
Cairo90–26030–2201999–2002
Dakar30–602008–2009
Bangkok40–90 (−)1995–2008
Beijing150–180 (−)95–1551999–2008
Delhi50–30050–2502004–2009
Dhaka>100 (+)>30 (+)2002–2006
Hong Kong40–50 (−)20–401998–2008
Jakarta60–100 (−)2001–2007
Manila40–5020–30 (−)2001–2008
Seoul60–80 (−)1995–2007
Shanghai90–110 (−)2002–2007
Tokyo15–30 (−)2001–2008
Tehran65–3702003
Santiago50–100 (−)20–30 (−)2000–2008
Sao Paulo40–70 (−)1996–2006
Los Angeles40–80 (−)1998–2008
Houston30–40 (−)1998–2008
New York City30–70 (−)1998–2008
Mexico City50–180 (−)20–25 (−)1990–2010
London20–351994–2004
Moscow35–50 (+)2006–2008
Milan35–60 (−)2000–2009
Istanbul45202002–2003

Hydroxyl radicals

Tropospheric self-cleaning capacity. An important role of UV radiation in the troposphere is the production of ˙OH radicals by photolysis of tropospheric ozone (Reaction 3 in Table 2) followed by reaction with H2O (Reaction 4, Table 2). The ˙OH radicals react with many of the gases emitted at the Earth's surface, including carbon monoxide (CO), methane (CH4) and other volatile organic compounds (VOCs), oxides of nitrogen and sulfur (NO2 and SO2), and hydrohalocarbons (HFCs and HCFCs). The reactions with ˙OH determine the atmospheric residence time of these gases, as well as their amount in the atmosphere since this is directly proportional to the product of emission rates and lifetime.

Understanding ˙OH is fundamental to understanding the chemistry of ozone and secondary aerosols as well. Cycling between ˙OH and HO2 (Reactions 5 and 6, Table 2) is essential for tropospheric ozone formation. Notably, ˙OH itself has a lifetime of only seconds, but it affects O3 on the time scale of hours to days, CO over months, and methane over a decade. For this reason, direct detection of ˙OH has focused on local short-term measurements, while longer-term impacts, for example, on methane lifetimes, have been estimated from global models.

Local measurements of ˙OH. Direct measurements of ˙OH are difficult because of its high chemical reactivity. Within seconds of being produced, ˙OH reacts with various gases (see previous paragraph) limiting its concentration to low values (about 106–107 mol cm−3 during daytime, smaller at night) that are exceedingly difficult to detect and quantify. The high reactivity also implies that ˙OH has high spatial and temporal fluctuations, being sensitive to variations in production (e.g., to variations in UV radiation, O3, and H2O) as well as loss (e.g., via reactions with CO, NOx, or VOCs). For this reason, it is important to note that locally measured ˙OH concentrations cannot be easily integrated spatially or temporally to estimate, for example, an annually averaged global ˙OH concentration. The main objective of direct local measurements is to evaluate whether the variations in ˙OH follow the expected variations in simultaneously measured meteorological (UV radiation, humidity) and chemical variables (O3, NOx, VOCs, etc.).

Several techniques have been developed over the past few decades to detect and measure concentrations of ˙OH. Measurements have been reviewed by Heard and Pilling81 and more recently by Stone et al.82 Most of the recent measurements are consistent with model calculations within a factor of approximately two, (e.g., urban locations including New York City,83 Tokyo,84 Mexico City,85,86 and Houston87). Studies in forested locations, specifically West Africa88 and northern Michigan89 also show reasonable agreement with models.

Measured ˙OH is much greater, by as much as an order of magnitude, than predicted by models in environments containing high concentrations of biogenic hydrocarbons (e.g., isoprene, methyl butenol, and terpenes) and low concentrations of NOx, including over the tropical forest of Suriname,90–92 Borneo,93,94 the Pearl River Delta (PRD),95,96 and suburban Beijing during low-NOx episodes.97 This apparent underestimation of ˙OH by models has led to a re-examination of the chemistry of isoprene at low NOx, and to the suggestion that at least part of the ˙OH initially lost by reaction with isoprene is later regenerated by secondary reactions.98,99 Simulations using an environmental smog chamber also indicate the need for some recycling of ˙OH by isoprene chemistry under very low NOx conditions, although not to an extent that would explain the large discrepancies between observations and models found over tropical Suriname (Fuchs et al., 2013).100

Several inter-comparisons between different ˙OH instruments show good agreement in some circumstances but also disagreement in others.82,101–103 The largest discrepancies appear to occur in environments dominated by biogenic hydrocarbons. For example, Mao et al.104 found a factor of two difference between laser-induced fluorescence and a chemical analysis methods at Blodget Forest, California.

Instruments have been developed recently to measure the total ˙OH reactivity, i.e., the rate at which ˙OH molecules are removed by reaction with the many constituents of sampled air (e.g., CO, VOCs, and NOx). The reactivity of ˙OH provides an important constraint on the budget of ˙OH since it must essentially balance the rate of production. Measurements show that this reactivity is larger than predicted from the simple sum of typically known constituents, indicating the presence of substantial but unmeasured amounts of other reactive compounds, by values ranging from 25–35% in Tokyo,105 about 30% in a terpene-rich mid-latitude forest,106 a factor of 2 in the Pearl River Delta107 and 60–90% in a boreal forest.108,109 The missing compounds are presumed to be a multitude of partly oxygenated organic compounds (aldehydes, ketones, etc.) formed during the ˙OH-initiated photo-degradation of VOCs.

Despite unresolved differences among various ˙OH instruments and models, some fundamental aspects of the photo-chemistry have been clearly demonstrated. Important in the present context is the theoretical expectation that, for relatively clean conditions, concentrations of ˙OH should scale more or less linearly with the photolysis of O3 to generate O* (Reaction 3, Table 2), j(O3), which in turn is dependent on the amount of UV radiation. This linear correlation has now been re-confirmed by direct measurements of ˙OH and j(O3) over a year at Mace Head, Ireland110 and is in agreement with earlier observations in the tropical Atlantic111 and in the European Alps.112 Because tropospheric j(O3) values are sensitive to the overhead ozone column, with a ∼1.5% increase in j(O3) for each 1% decrease in the O3 column (see Table 1 of McKenzie et al.,37), these studies reaffirm the importance of stratospheric ozone to tropospheric ˙OH and to the photochemistry of the lower atmosphere.

Global ˙OH models. Estimates of long-term changes in global ˙OH are uncertain and variable. Empirical estimates, based on the measured concentrations of trace gases, such as methyl chloroform, whose emissions and ˙OH kinetics are well known, are difficult due to large changes in emissions of suitable gases. An analysis of the decline in concentrations of methyl chloroform by Montzka et al.113 concluded that globally averaged ˙OH varied by less than ±5% during 1997–2007. On the other hand, Monteil et al.114 interpreted measurements of methane isotopes (13C) to infer that the slowing of CH4 trends in the early 2000s was due to increasing concentrations of ˙OH, at about 5% per decade, due to global increases in NOx emissions. Thus, the direction and magnitude of recent trends in global ˙OH remains unclear.

Estimates over longer time scales are largely based on models, which differ significantly. This was demonstrated by the recent intercomparison of 16 global chemistry-transport models for predictions of the methane lifetime, which is limited by reaction with ˙OH.5,7 The modelled mean lifetime of methane was 8.6 ± 1.2 years (range 6.4–11.6 years) for the year 2000. Pre-industrial (1850) to present day (2000) changes in ˙OH were either positive or negative (see Fig. 6), depending largely on how each model specified relative changes in emissions of CO and NOx. Pike and Young115 showed that global concentrations of ˙OH (and therefore the lifetime of CH4) were sensitive to how models represent ˙OH recycling by isoprene, which remains uncertain, as discussed above. If such buffering of ˙OH by biogenic VOCs is pervasive, it casts doubt on the strong sensitivities to anthropogenic emissions of CO and NOx shown in Fig. 6 for current models.


Fig. 6 Changes in globally averaged hydroxyl radicals (˙OH) between pre-industrial times (1850) and present day (2000) calculated by 16 different models (model-mean for the year 2000 ˙OH ∼1.1x106 molec cm−3) for relative changes in emissions of carbon monoxide (ΔCO) and nitrogen oxides (ΔNOx) specified within each model (From Naik et al.5).

An alternate approach to estimating changes in ˙OH based on sulfate isotopic studies suggested a 10% decrease in global ˙OH since pre-industrial times,116 broadly in line with some of the models reported by Naik et al.5 However, the accuracy of this method remains untested.

The multi-model mean changes in predicted ˙OH concentrations at the surface for the year 2100 and two different emission scenarios, RCP2.6 and RCP8.5, have been calculated7 and are shown in Fig. 7. Substantial reductions in ˙OH are expected throughout much of the southern hemisphere due to large increases in methane (CH4) in the RCP 8.5 scenario, while regional increases and decreases occur in both scenarios due to changes in shorter-lived precursors NOx and CO. Another model study focused on the recovery of stratospheric O3 to 1980 levels (holding all other factors constant) predicted that global concentrations of ˙OH will decrease by 1.7% due to the lower tropospheric UV radiation levels.53


Fig. 7 (a) Annual average surface ˙OH concentration, mean of 14 models, for the year 2000. (b) Model-mean % change in surface ˙OH concentrations in 2100 relative to 2000 for the IPCC RCP 2.6 emission scenario; (c) same for RCP 8.5 scenarios (From Voulgarakis et al.7).

Global models are also sensitive to climate change, including changes in temperature, humidity, stratospheric ozone, and uncertain NOx emissions from natural sources such as biomass burning and lightning.117–119 Thus, multi-model averages such as those shown in Fig. 7 do not truly reflect the model variability or actual uncertainties. However, in clean marine atmospheres, concentrations of ˙OH are well predicted by models,120,121 but other unidentified oxidants appear to be important.120 These other oxidants could be halogens.

Climate-mediated changes in air quality

Air pollution is a complex, multifaceted problem that can only be correctly considered when integrated within the whole Earth system. Direct emissions from human activities are well recognized, but emissions that would otherwise be considered natural can also change due to, for example, deforestation, biomass burning, and even feedback between air quality, climate change (especially the hydrological cycle), and ecosystem health. Atmospheric transport of pollutants and their precursors is subject to circulation patterns that are likely to change under a changing climate. In particular, changes in the frequency of stagnation episodes that limit the dispersion of pollutants may have large impacts on air quality in affected areas. Chemical transformations, e.g., those making ground-level O3 from the photo-oxidation of hydrocarbons and nitrogen oxides, are sensitive to climate variables including temperature and moisture, as well as UV radiation. Removal of pollutants occurs mainly via contact with the Earth's surfaces (dry deposition) or scavenging by precipitation (wet deposition). Both could change significantly in the future, e.g., changes in land-use altering rates of dry deposition, and changes in precipitation patterns modifying wet deposition rates.

Several potentially major interactions are discussed here. Many other feedback processes are possible and may be plausible but are not fully understood and cannot yet be quantified reliably. These are, by and large, outside the scope of the present assessment, except to the extent that we recognise their existence and therefore provide a cautionary note that the general aspects of our assessment need to be evaluated carefully for any given location with full consideration of these additional factors.

Ozone. Concentrations of ozone in urban environments are determined by a number of key factors. Firstly, the amount of ozone in the air entering the urban environment may be important. Then, within the urban airshed, reactions involving a range of emitted chemicals, most notably VOCs and NOx, produce ozone as a result of UV-driven photochemistry. These chemicals arise from both anthropogenic and biogenic sources, the latter often being outside normal air quality management. Finally, a number of processes such as dry deposition (loss at the surface) might remove ozone from the atmosphere.

Background ozone concentrations can be influenced by long-range air transport, and this is likely to be an important factor in the future, transporting air between continents.122 The background ozone concentrations can also be altered by changes in stratosphere/troposphere exchange and by the changes in global atmospheric composition, notably some greenhouse gases.119 By analysing the natural variability of stratospheric transport, it has been estimated that changes in stratospheric circulation due to climate change will lead to around a 2% increase in tropospheric ozone in the northern mid-latitudes by the end of this century.123 Although small, the significant geographic extent implies that this might be an additional factor affecting air quality.

Increased temperatures at ground level are expected to increase biogenic emissions of reactive organics (e.g., ref. 124 and 125). However, emissions may well depend on other factors as well, such as water stress on the plants.32 Indeed, changes in the climate, coupled to decreases in air quality, can substantially alter biogenic activity in ways that are difficult to predict.

Models suggest that other critical processes are also likely to be altered by climate change. Variations in cloudiness can alter the rate of photochemical production of ozone. Increased surface heating can result in changes in atmospheric movement (wind speed, both horizontal and vertical). Cloudiness can also be altered by human activity, with evidence that controls on air pollution have increased solar radiation at some locations.126 Finally, changes in rainfall patterns and cloudiness can alter the rate of removal of both reactive precursors and ozone itself.26 Estimates predict increasing ozone concentrations at ground level throughout the 21st century, driven by all of these meteorological factors, which are regionally dependent. This can be offset by changes in emissions from human activity, which may either augment or reverse the overall trend, depending on the levels of controls implemented.26,127 Significant regional air quality changes may result, even if only episodically.128

Particles. The impact of climate change on aerosols remains highly uncertain. As mentioned above, increasing temperatures will increase biogenic gas emissions. The oxidation of these compounds will produce aerosols. Additional wild fires could become more important as a source of particulate matter.129 These aerosols can scatter radiation and reduce warming (a negative feedback), and also impact upon cloud properties.130 Changes in clouds in turn can alter the transformation (growth, chemistry) of aerosols in the atmosphere. In polluted environments the changes induced by climate will be overwhelmed by anthropogenic emissions, but in locations where anthropogenic emissions are small these changes could be significant. However, the net effect of climate change on aerosols remains unclear.127,129 While there has been a lot of work in this area (e.g., on climate/aerosol feedback models131) the level of scientific understanding remains very low.

Biological interactions between air-pollutants and climate change. Increased ambient temperatures may interact directly or indirectly to exacerbate the effects of pollutants such as O3 in humans. A study of cardiovascular and respiratory mortality in 2002 to 2006 in Buenos Aires showed a relative risk of 1.0184 (95% confidence interval (CI) = 1.0139–1.0229) on the same day for each 1 °C increase in temperature.132 Another study in several large cities in the UK reported that temperature increased mortality from cardiovascular and respiratory diseases and other non-accidental causes.133 These authors also reported that the mean mortality rate ratio for O3 was 1.003 (95% CI = 1.001–1.005) per 10 μg m−3 increase in concentration. On hot days (greater than the whole-year 95th centile) this increased to 1.006 (95% CI = 1.002–1.009) but was only statistically significant for London. A study in older men in 2000 to 2008 in Boston showed that greater ambient temperature was associated with decreases in heart rate variability via dysfunction of the autonomic nervous system.134 These warm-season associations were significantly greater when ambient ozone concentrations were above the median but were not affected by particulates (PM2.5). These studies are consistent with earlier analyses of the 2003 European heat wave episode135,136 which attributed a significant fraction (20–60%) of excess mortality to the effects of elevated levels of O3 and PM.

Change in climate also may affect human health indirectly. A study on allergic respiratory diseases and bronchial asthma showed that while exacerbation is related to air-pollutants, amounts of allergen in the air are also important.137 The presence of allergenic pollens in the atmosphere might be prolonged by climate change and increase frequency and severity of these diseases.

Halogenated organic and other substitutes in the troposphere

Toxicity and risks of replacements for ozone depleting chemicals to humans and the environment. The United States Environmental Protection Agency (USEPA) has a regulatory process for evaluating alternatives for Ozone Depleting Substances (ODS) prior to their wide-spread use in the U.S. Anyone planning to market or produce a new substitute must provide 90 days’ advance notice to the Significant New Alternatives Policy (SNAP) program at USEPA of their intent as well as providing health and safety information before introducing it into interstate commerce in the U.S. Normally the health and safety information will include information on chemical and physical properties, flammability and basic toxicological information, and more recently, global warming potential. The SNAP program reviews the information in the context of the proposed use, and issues one of 4 decisions: acceptable; acceptable subject to use conditions; acceptable subject to narrowed use limits; and unacceptable. This information on a particular compound is continually updated so that compounds may be proposed for additional uses or additional information may be added to the portfolio for a particular use that could change the initial decision.

Updates on selected halocarbons.
Brominated substances. Natural bromo-carbons bromoform and dibromomethane are emitted from the oceans, and their emission strengths and role in the atmosphere are becoming better understood. These compounds release bromine upon oxidation in the atmosphere that is generally observed as bromine monoxide (BrO). The presence of bromine can lead to depletion of ground level ozone.138 However, observations in other marine locations do not find such events occurring.139