Indoor air pollution is ranked as one of the world's greatest public health risks (Wolverton, 1997). The United Nations Development Program estimated in 1998 that over 2 million humans die each year due to the persistence of deleterious indoor air (Brennan and Withgott, 2005). It has also been estimated that globally 14 times as many deaths occur from poor indoor air quality compared with ambient air pollution (Brennan and Withgott, 2005).
Because humans in industrialized countries spend about of 80% to 90% of their time indoors (Orwell et al., 2004; Wolverton, 1997), negative societal consequences due to polluted indoor air can be great (Fisk and Rosenfeld, 1997). For example, the cost of unhealthy indoor air in Australia has been estimated at $12 billion annually due to losses in productivity, higher medical costs, more absenteeism, and lower earnings (Wood, 2003). In addition, the health burden associated with indoor air pollution does not appear to be equal for developing and developed countries. Australia's Commonwealth Science Council has suggested that 9 of 10 deaths due to indoor air are experienced by the developing world (Brennan and Withgott, 2005).
Ozone (O3), a photochemical oxidant with a redox potential of +2.07 V, is one of the most powerful naturally occurring oxidants (Maroni et al., 1995; Mustafa, 1990). It is considered a secondary ambient pollutant and one of the components of tropospheric smog, which can adversely affect human health and property (Mustafa, 1990). Ozone is produced within the troposphere by chemical reactions involving free radicals. It is formed during the reaction of carbon monoxide (CO) and volatile organic compounds (VOCs) in the presence of nitrogen oxides (NOX), oxygen and sunlight. There is strong evidence to suggest that average ozone concentrations have been rising during the past century as a result of increased input of the precursors of ozone into the atmosphere (Mustafa, 1990). Automobiles are the principal contributors to secondary tropospheric ozone generation (Maroni et al., 1995).
Ozone as an indoor air pollutant can be prevalent in homes and offices due to infiltration of outdoor ambient air indoors (Weschler, 2000). Ozone-emitting equipment such as copy machines, laser printers, ultraviolet lighting, and some electrostatic air purification systems may also contribute to indoor ozone levels (Maroni et al., 1995; Weschler, 2000). Ozone generation from appliances such as photocopiers on average yield 5.2 mg·h−1 and laser printers on average produce 1.2 mg·h−1; however, concentrations could vary based on equipment maintenance (Black and Worthan, 1999; Weschler, 2000). Depending on the air exchange rates between outdoor and indoor environments, indoor air has been reported to contain from 10% to 50% of outdoor values (Weschler, 2006) to five to seven times the contaminant concentrations of ambient urban air (Brown et al., 1994; Orwell et al., 2004).
The principal groups at risk to ozone toxicity are organisms in which the primary route of exposure is via inhalation (Wright and Welbourn, 2002). Daily inhalation of indoor ozone for these organisms (such as humans) are estimated to be between 25% and 60% of total ozone intake (Weschler, 2006). Major toxic effects of ozone in humans include alterations in pulmonary function in addition to cellular and biochemical endpoints (Wright and Welbourn, 2002). Exposure to ozone also may result in pulmonary edema, hemorrhage, inflammation, and extensive lesions on the lung tissue, trachea, and upper bronchi (Mehlman and Borek, 1987).
Indoor exposures to ozone are often accompanied by exposure to the products of ozone-initiated oxidative reactions (such as isoprenes, styrenes, terpenes, sesquiterpenes, and unsaturated fatty acids) (Weschler, 2004). The average daily intakes of ozone oxidative products are roughly one-third to twice the average daily intake of ozone alone (Brown et al., 1994). Some of the oxidative products of ozone that are known or suspected to adversely affect human health include formaldehyde, acrolein, hydrogen peroxides, and fine and ultrafine particles (Weschler, 2004). Common indoor sources of reactive chemicals to ozone include occupants themselves, soft woods, linoleums, certain paints, polishes, cleaning products, soiled fabrics, and soiled ventilation filters. Indirect evidence supports connections between human morbidity/mortality and exposures to indoor ozone and its oxidative products (Weschler, 2006).
As indoor air pollution poses concerns for human health, cost-effective and easy-to-implement methods are needed to eliminate or reduce concentrations. Activated charcoal filters reduce air pollutants but installation and maintenance costs can be high (Wolverton, 1997). As an alternative, foliage plants could be used to sequester some types of air pollution, although their effectiveness appears to be plant species specific and air pollutant dependent (Wolverton, 1986).
Wolverton (1984) evaluated common foliage plants in controlled chambers (that simulated indoor environments) for their ability to reduce concentrations of several air pollutants. Depletion rates of known concentrations of air pollutants within the enclosed chambers containing plants were measured. Wolverton concluded that of the taxa selected, common spider plant and golden pothos most effectively reduced various air pollution concentrations (e.g., formaldehyde, nitrogen dioxide, and carbon monoxide) from closed chambers (Wolverton, 1984, 1986). Because ozone was not one of the indoor air pollutants evaluated in the previous study (Wolverton, 1984) and it is an important indoor air pollutant for which mitigation methods are limited, the objective of the current study was to determine the effectiveness of using common foliage houseplants in reducing concentrations of ozone within a simulated indoor environment.
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