13.2. Some effects on the environment

13.2.1. Smog

The word “smog” was originally generated by the combination of words “smoke” and “fog” to describe a type of air pollution that refer a mixture of natural fog and several air pollutants. Today, it is a common term applied to different air pollution events mainly formed in large cities.

“London type smog” or sulphurous smog was frequently formed from the 19th century to the mid 20th century in industrialized regions. One of the most disastrous smog event happened in 1952, in London, when thousand of deaths were associated with air pollution. This type of smog is a consequence of large amount of coal burning in cities or industrial areas. Under stable atmospheric stratification, sulphur dioxide, sulphuric acid, soot particles and other pollutants produced from the fossil fuel sources can be piled up. Cold, damp, foggy weather favours the formation of smog. London type smog was largely eliminated during the 20th century, by sulphur dioxide and smoke emission reduction strategies. However, similar air pollution events can occurs nowadays in winter season (see Figure 13.8) due to the large emission of particulate matter, nitrogen oxides and other pollutants from traffic and residential heating combined with a long-lasting high-pressure system.

Another type of smog is the “Los Angeles type smog” or photochemical smog. This type of air pollution was first identified in 1954 in Los Angeles (Haagen-Smit and Fox, 1954). The photochemical smog consists of ozone and other closely related secondary pollutants that are produced by photochemical reactions from precursor compounds, which are emitted to the atmosphere mainly by transport. However, the photochemical smog formation is also influenced by the weather conditions. Ozone formation is driven by sunlight. Severe ozone events are often associated with persistent high-pressure systems (anticyclones), light wind and subsidence inversions, which limit vertical dispersion of pollutants. The chemistry of ozone formation is also depends on the temperature. Higher temperature accelerates the ozone formation. The elevated ozone levels are always associated with temperatures in excess of 20°C (Sillman, 2003). Ozone formation typically requires a few hours and during this time, the air mass from the city can be transported away. Therefore, high ozone values generally are found downwind of major cities rather than in the downtown. However, occasionally, most severe photochemical smog can be formed in city centre in case of very light wind. Geographical location of the city (if the city is situated for example in a valley) can facilitate the formation of elevated ozone concentrations.

13.2.2. Acid rain

Acid rain is a commonly used term of acidic deposition. During this process, high amounts of stong acids (such as nitric and sulfuric acids) and other acid-forming substances (ions, gases and particles) are removed from the atmosphere to the surface by both dry and wet deposition.

Compounds of acidic deposition are formed from the gaseous emission of sulphur dioxide (SO2), nitrogen oxides (NOx) and ammonia (NH3) and particulate emission of acidifying and neutralizing compounds. Due to the long range transport of these acidic compounds, the harmful effects of acid deposition could be observable even more hundreds of kilometers from the emission sources. Acidic deposition caused a critical environmental stess that affects soils, forested lanscapes (Figure 13.9) and aquatic ecosysyems in the last quatrer of 20th century in North america, Europe, and Asia (Driscoll et al., 2001). However, due to the pronounced decrease in sulphur dioxide emission, and based on the continuous pH measurements of precipitation, the acidity of precipitation had decreased significantly in North America and Europe from the 1970s. At the same time, because of losses in soil buffering, the forest ecosystem is currently much more sensitive to acid rain inputs than previously predicted (Likens, 2013).

Harmful effect of acid deposition on forest trees

Figure 13.9: Harmful effect of acid deposition on forest trees

13.3.3. Crop and forest damage by ozone

The background surface ozone concentration increased steadily worldwide in the last decades of 20th century (Vingarzan, 2004). This explicit trend may slow in previous years, but measured values are still high in Europe (EEA, 2011). Moreover, based on air quality model simulations, significant rise of surface ozone concentration is predicted for the future (see e.g. Meleux et al., 2007) causing several environmental damages. Ozone in the near surface layer is one of the most important phytotoxic air pollutants that can cause injury to plant tissues, reduction in plant growth and productivity via ozone uptake, especially through the stomata. Ozone can also affect the mechanism of CO2 exchange between vegetation and the atmosphere (Sitch et al. 2007). Several studies have reported that ozone can cause the most damage to forest vegetation. Impaired trees biomass growth caused by the elevated ozone concentration.

To estimate the effects of ozone on vegetation, concentration based metrics (e.g. AOTX – accumulated O3 exposure over a threshold X value) have been proposed at the end of 20th century. However, from the biological aspect, the response of vegetation to ozone is more closely related to the absorbed dose through the stomata than to external ozone exposure (e.g. Musselmann et al. 2006). To characterize the vegetation damage caused by the ozone, in the past decade, flux-based ozone exposure metrics have been favoured as opposed to concentration-based indices. The differences between AOTX and O3 flux indices are more considerable under dry climatic conditions where vapour pressure deficit and soil moisture deficit limit the stomatal conductance.

The ozone exposure can be estimated by more or less sophisticated deposition models for several types of vegetation or for a region (see. e.g. Mészáros et al., 2009). In such models, the ozone flux is controlled by ozone concentration and by deposition velocity via parameterization of the canopy and stomatal conductances. In general, in the models a multiplicative algorithm of stomatal conductance is applied (see Chapter 12). This method includes functions for the effects of photosynthetically active radiation, air temperature, soil water content, and other parameters affecting the stomatal conductance. Plant stomatal conductance and calculation of the deposition velocity play a key role in most deposition models applied for risk assessment and for estimation climatic effects of tropospheric ozone.

The total ozone flux is the product of the deposition velocity and ozone concentration. Therefore, on the distribution of the flux, the effects of both concentration and deposition fields are apparent. In case of high concentration, generally high flux can be observable. However, the predicted total ozone flux can be relatively low due to low deposition velocities or relatively high in case of high deposition velocity in some regions, demonstrating possible flaws in the assumption that damage can be directly related to ozone concentrations (See Figure 13.10). The plant response and therefore the effective ozone load are more closely related to the ozone flux than to the atmospheric concentrations. The weather situation and soil properties through the stress effects on plants can retard the deposition. Therefore, in some cases lower amounts of ozone can be settled from the atmosphere, even if the ozone concentration is elevated.

Estimated AOT40 and total flux of ozone over deciduous forest in Hungary in July 1998

Figure 13.10: Estimated AOT40 and total flux of ozone over deciduous forest in Hungary in July 1998. Model simulations were carried out by TREX transport-deposition model (Eötvös Loránd University). Meteorological data were obtained from ALADIN numerical weather prediction model used at Hungarian Meteorological Service.


Andersen Z.J., Kristiansen L.C., Andersen K.K., Olsen T.S., Hvidberg M., Jensen S.S., Ketzel M., Loft S., Sørensen M., Tjønneland A., Overvad K., and Raaschou-Nielsen O.. 2012. Stroke and Long-Term Exposure to Outdoor Air Pollution From Nitrogen Dioxide : Cohort Study. Stroke. Vo. 43. 320-325.

Bell M.L. and Davis D.L.. 2001. Reassessment of the Lethal London Fog of 1952: Novel Indicators of Acute and Chronic Consequences of Acute Exposure to Air Pollution. Environmental Health Perspectives. Vo. 109. 389–394.

Driscoll C.T., Lawrence G.B., Bulger A.J., Butler T.J., Cronan C.S., Eagar C., Lambert K.F., Likens G.E., Stoddard J.L., and Weathers K.C.. 2001. Acidic deposition in the Northeastern United States: sources and inputs, ecosystem effects, and management strategies. BioScience. Vo. 51. No. 3.. 180-198.

EEA Technical Report, 2011: Air pollution by ozone across Europe during summer 2010. EEA Report. No 6/2011. ISBN 978-92-9213-210-1.

Emberson L.. 2003. Air pollution impacts on crops and forests: An introduction. In: Emberson, L., Ashmore, M., Murray, F., eds, Air pollution impacts on crops and forests. A global assessment. Imperial College Press, London. Vo. 4. 3-29 pp.

Haagen-Smit A. J. and Fox M.M.. 1954. Photochemical ozone formation with hydrocarbons and automobile exhaust. Air Repair. Vo. 4. 105–109.

Helfand W.H., Lazarus J., and Theerman P.. 2001. Donora, Pennsylvania: An environmental disaster of the 20th century. American Journal of Public Health. Vo. 91. No. 4. 553.

Likens G.E.. 2013. Acid rain. In: Weathers, K.C., Strayer, D.L. Likens, G.E. (eds): Fundamentals of Ecosystem Science. Academis Press. Section V., Chapter 15. 259–264. ISBN: 978-0-12-088774-3.

Meleux F., Solmon F., and Giorgi F.. 2007. Increase in summer European ozone amounts due to climate change. Atmospheric Environment. Vo. 41. 7577–7587.

Mészáros R., Zsély I.Gy., Szinyei D., Vincze Cs., and Lagzi I.. 2009. Sensitivity analysis of an ozone deposition model. Atmospheric Environment. Vo. 43. 663–672.

Musselman R.C., Lefohn A.S., Massman W.J., and Heath R.L.. 2006. A critical review and analysis of the use of exposure- and flux-based ozone indicies for predicting vegetation effects. Atmospheric Environment. Vo. 40. 1869–1888.

Nemery B., Hoet P.H.M., and Nemmar A.. 2001. The Meuse Valley fog of 1930: an air pollution disaster. Lancet. Vo. 357. No. 9257. 704–708.

Pöschl U.. 2005. Atmospheric Aerosols: Composition, Transformation, Climate and Health Effects. Angewandte Chemie International Edition. Vo. 44. 7520–7540.

Raub J.A. and Benignus V.A.. 2002. Carbon dioxide and the nervous system. Review. Neuroscience and Biobehavioral Reviews. Vo. 26. 925–940.

Sillman S.. 2003. Tropospheric Ozone and Photochemical Smog. in: Holland, H.D., Turekian, K.K., eds: Treatise on Geochemistry. Environmental Geochemistry, Elsevier. Volume 9. 407–431. ISBN: 978-0-08-043751-4.

Sitch S., Cox P.M., Collins W.J., and Huntingford C.. 2007. Indirect radiative forcing of climate change through ozone effects on the land-carbon sink. Nature. Vo. 448. 791–794.

Vingarzan R.. 2004. A review of surface ozone background levels and trends. Atmospheric Environment. Vo. 38. 3431–3442.

Weschler C.J.. 2006. Ozone’s Impact on Public Health: Contributions from Indoor Exposures to Ozone and Products of Ozone-Initiated Chemistry. Environmental Health Perspectives. Vo. 114. 1489–1496.