Nanometer-sized emissions from municipal waste incinerators: A qualitative risk assessment

Journal of Hazardous Materials 320 (2016) 67–79

Contents lists available at ScienceDirect

Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Review

Nanometer-sized emissions from municipal waste incinerators: A qualitative risk assessment David R. Johnson GHD, 1755 Wittington Place, Suite 500, Dallas, TX 75234, USA

h i g h l i g h t s

g r a p h i c a l

a b s t r a c t

• Incinerators generate particles in the nanometer-scale range (INPMWI ).

• INPMWI exposure in ambient atmosphere is low compared to other sources. • INPMWI toxicity may be moderate due to uncertainties in particle composition. • A qualitative risk assessment of INPMWI predicts low-moderate risk from INPMWI . • Risk from INPMWI from incinerators appears less than initially perceived.

a r t i c l e

i n f o

Article history: Received 23 May 2016 Received in revised form 29 July 2016 Accepted 5 August 2016 Available online 5 August 2016 Keywords: Incidental nanoparticles Municipal waste incinerators Emissions Particulate matter Risk

a b s t r a c t Municipal waste incinerators (MWI) are beneficial alternatives to landfills for waste management. A recent constituent of concern in emissions from these facilities is incidental nanometer-sized particles (INPMWI ), i.e., particles smaller than 1 micrometer in size that may deposit in the deepest parts of the lungs, cross into the bloodstream, and affect different regions of the body. With limited data, the public may fear INPMWI due to uncertainty, which may affect public acceptance, regulatory permitting, and the increased lowering of air quality standards. Despite limited data, a qualitative risk assessment paradigm can be applied to determine the relative risk due to INPMWI emissions. This review compiles existing data on nanometer-sized particle generation by MWIs, emissions control technologies used at MWIs, emission releases into the atmosphere, human population exposure, and adverse health effects of nanometer-sized particles to generate a qualitative risk assessment and identify data gaps. The qualitative risk assessment conservatively concludes that INPMWI pose a low to moderate risk to individuals, primarily due to the lack of relevant toxicological data on INPMWI mixtures in ambient particulate matter. © 2016 Elsevier B.V. All rights reserved.

Abbreviations: CB-MWI, combustion-based municipal waste incinerator; ENP, engineered nanoparticle; GB-MWI, gasification-based municipal waste incinerator; IGCC, integrated gasification combined cycle; INP, incidental nanometer-sized particle; INPMWI , incidental nanoparticles from municipal waste incinerators; LCA, life cycle assessment; MWI, municipal waste incinerator; NAAQS, national ambient air quality standards; NNP, natural nanometer-sized particles; NO, nitric oxide; PM, particulate matter; PMx , particulate matter at or below the designated size (in micrometers); USEPA, United States Environmental Protection Agency. E-mail addresses: [email protected], [email protected] http://dx.doi.org/10.1016/j.jhazmat.2016.08.016 0304-3894/© 2016 Elsevier B.V. All rights reserved.

68

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

Contents 1. 2.

3.

4.

5.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68 Incidental nanometer-sized particle characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 2.1. Nanometer-sized particle classification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 2.2. Municipal waste incinerators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 2.3. Synthesis of incidental nanometer-sized particles by combustion-based municipal waste incinerators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 2.4. Fate of engineered nanomaterials in combustion-based municipal waste incinerators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 2.5. Predicted INP synthesis in gasification based municipal waste incinerators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69 Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71 3.1. Flue gas and emissions control technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71 3.2. Incidental nanometer-sized particle release into the atmosphere . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72 3.3. Nanometer-sized particle exposure in personal microenvironments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73 Toxicity of nanometer-sized particles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 4.1. Toxicity of incidental nanometer-sized particles in animal models and human volunteers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 4.1.1. Effects of nanometer-sized particles on the respiratory system. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .74 4.1.2. Effects of nanometer-sized particles on the cardiovascular system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76 Human health risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76 5.1. Qualitative risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76 5.2. Data gaps . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77 Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78

1. Introduction Municipal waste incineration is an increasingly beneficial alternative to landfills for waste management. The benefits range from reduced landfill space to reduced greenhouse gas production (e.g., methane) [1]. The process consists of combusting organic substances from municipal waste under conditions that generate heat, which can then be used for energy production. The byproducts from this process include gases (e.g., nitrates, sulfates, and carbon dioxide), water, and coarse and fine particulates (i.e., bottom ash/soot and fly ash, respectively). Gases and particulates pass through emission control technologies to remove solid and gaseous contaminants before flue gas is released into the ambient atmosphere. Emission control technologies for municipal waste incinerators (MWI) have continually improved over the years, with high efficiency removal of emissions from flue gas [2]. Yet the public still has deep concerns about potential health effects from emissions that escape the emission control technologies [3–5]. These concerns stem from the historical issues, known chemicals released in emissions, and uncertainty of what unknown chemicals are also being released in the emissions. Recent public concern surrounds the release of particulate matter (PM) with mean aerodynamic diameters below 2.5 ␮m (i.e., PM2.5 ), a subset of total PM and a primary pollutant currently regulated under the National Ambient Air quality Standards (NAAQS) administered by the U.S. Environmental Protection Agency (USEPA). Epidemiological studies demonstrate positive associations between long-term exposure to PM2.5 concentrations and mortality, cardiovascular effects, respiratory effects, and reproductive and developmental effects, as well as associations between short-term exposure to PM2.5 and mortality and morbidity health endpoints [6,7]. In order to protect human health, the USEPA continues to reassess the ambient air quality standard for PM2.5 , most recently resetting the standard at 12 ␮g/m3 [8]. However, increased concern has surrounded PM2.5 particles below the micrometer range, i.e., particles that are within the 1–999 nm size range (Fig. 1). These nanometer-sized particles have very little mass compared to micrometer-sized particles, but are higher in particle number count. These particles also have a higher surface area for reactivity (e.g. aluminum) [9]. Smaller nanometersized particles (e.g., ultrafine particles or nanoparticles) have the potential to deposit deep into the lungs where the gas-blood interchange occurs. These particles can potentially affect lung function or cross into the bloodstream and affect peripheral regions of the

human body. Scientists are conducting toxicology studies to better understand the uptake and effects of inhaled nanometer-sized particles in the body, however, it is highly plausible that citizens may make incorrect comments about the dangers of nanometersized particle exposure, uptake, and toxicity to human health. These comments and opinions may be a result of inaccurate information sources (e.g., internet and social media), as well as selective use and unintended misinterpretations of scientific literature. Regardless of the source, it is highly likely that the result will be fear of the uncertainty. This uncertainty may result in extra precaution when it may not be warranted or necessary, such as during regulatory permitting of MWIs, especially since there currently are no national standards for ambient atmospheric concentrations or particle numbers for ultrafine particles (i.e., nanometer-sized particles). Risk assessments can be utilized as a way of addressing uncertainties—perceived and actual—associated with nanometersized particle emissions from MWIs. The risk assessment paradigm [10] characterizes site conditions, exposure, effects, and risk associated with physical and chemical stressors. When sufficient data are available, a quantitative risk assessment can generate a measurable threshold that is defined as a chemical concentration below which undue risk is not anticipated to occur. In contrast, a qualitative risk assessment is conducted when there is insufficient data to generate a measurable threshold. Instead, a qualitative risk assessment uses strong reasoning and a weight of evidence approach from existing data to support a conclusion that the exposure scenario for a chemical of concern is not likely to cause adverse health effects [11–13]. As it stands, stakeholders might posit that there are limited data available on exposure and health/environmental effects of nanometer-sized particles and thus there is a high degree of risk. However, it is possible to use existing peer-reviewed literature to glean insight into potential exposure and effects of nanometersized particles so as to qualitatively reduce the perceived risk level to a more realistic and accurate risk level [11]. Thus, this review utilizes existing peer-reviewed literature to synthesize a qualitative risk assessment of nanometer-sized particles released from MWIs. The risk assessment will consist of nanometer-sized particle characterization (e.g., definitions, synthesis schemes, and physicochemical properties) (Section 2), exposure (emissions control technologies, release from MWI) (Section 3), effects (laboratory animal and human volunteer studies, epidemiology) (Section 4), and the qualitative risk assessment (Section 5). This review also

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

identifies data gaps (Section 6) that need to be addressed in order to strengthen risk assessments for nanometer-sized particles from MWIs and allow subsequent risk assessments to transition from qualitative to quantitative. 2. Incidental nanometer-sized particle characterization 2.1. Nanometer-sized particle classification Nanometer-sized particles are generally divided into three categories: natural, engineered, and incidental. Natural nanometersized particles (NNP) exist in nature in a variety of forms, such as clays and metals in soils, gases and metal particulates from volcano eruptions, and PM from natural fires. NNPs are a critical component of the environment, giving environmental matrices their unique characteristics. Several NNPs are also dynamic materials, often changing size and composition depending upon the environmental conditions [14,15]. NNPs are generally not monodispersed but agglomerates and/or aggregates with other NNPs in the environmental matrix (e.g., humic acids in soils). Humans are exposed to NNPs through normal environmental processes (e.g., breathing dust) and activities (e.g., working and recreating in soils and natural waters). Engineered nanoparticles (ENP) are nanoparticles that are specifically designed by scientists. ENPs can be made with a wide variety of chemicals, both inorganic (i.e., metals) and organic (i.e., carbon based) elements, using different synthesis techniques (e.g., milling, plasma synthesis, vapor deposition, wet lab synthesis). By changing the synthesis conditions and settings, scientists can fine-tune the size and composition of ENPs to enhance particle properties. Products containing ENPs range in the thousands, with applications spanning electronics and biotechnology to paints and recreational equipment [16,17]. ENPs will continue to be incorporated in a wide variety of consumer products at a rapid pace and are predicted to have an economic value of over $4 trillion by 2018 [18]. Incidental nanometer-sized particles (INP) are unintentional by-products of anthropogenic processes or activities, e.g., fossil fuel combustion, welding, mining, and grinding. Examples of INPs are soot, fly ash, smoke, PM, and metal particulates, to name a few. INPs come from either primary or secondary sources. Primary sources (e.g., fire) generate INPs directly, whereas secondary sources (e.g., factories, vehicles, construction sites) generate gases that can condense and form INPs. This manuscript will focus primarily on primary INPs generated from municipal waste incinerators (INPMWI ). 2.2. Municipal waste incinerators MWIs, also known as waste-to-energy incinerators, recover energy from municipal waste while reducing landfill growth. MWIs are either combustion-based incinerators (CB-MWI) or gasification-based incinerators (GB-MWI). CB-MWIs add municipal solid waste (or fuel) directly to the combustion chamber, then burns the waste at high temperature and high oxygen conditions to generate heat and carbon dioxide. The waste can be moved by various mechanisms (e.g., grate, rotating kiln) to ensure more complete combustion and incineration. GB-MWIs use municipal solid waste (or feedstock) to generate “syngas” in a high temperature and low oxygen gasification chamber [19]. The syngas—consisting of methane, low molecular weight Cx Hx hydrocarbons, and carbon monoxide—can then be (1) combusted to generate heat or (2) used in chemical reactions to generate products, as is done for transportation fuels, chemicals, fertilizers, and substitute natural gas [19]. The resultant heat from both incineration processes is

69

used to heat steam for electric generation and residential heating [1]. Currently there are over 900 combustion-based incinerators in operation world-wide, with capacity ranging from 50 to 100 kiloton/year, whereas there are approximately 100 gasification incinerators worldwide, with capacity ranging from 10 to 250 kiloton/year [19,20]. 2.3. Synthesis of incidental nanometer-sized particles by combustion-based municipal waste incinerators The goal of incineration at MWIs is to generate as much energy from the organic components of municipal waste. The high temperature, high oxygen environment of a CB-MWI can result in high combustion efficiency (i.e., CO2 generation), but incomplete combusted organic material is also generated, some of which is small and easily aerosolizable (e.g., fly ash) and exits the combustion chamber with flue gas. The fate of the inorganic components of municipal waste is also affected by the incineration conditions. A metal can be vaporized if its boiling point is lower than the incinerator conditions. For instance, metals with lower boiling points (e.g., zinc and cadmium) are likely to be vaporized at lower temperatures, and metals with higher boiling points (e.g., cerium and titanium) are likely to be vaporized only at very high temperatures (i.e., >3000 ◦ C). Metal vapors follow the flue gas out of the incineration chamber; however, once the flue gas cools down, the metal vapors condensate and form nanometer-sized particles (i.e., INPMWI ). The size of these INPMWI is also dependent upon the boiling point of the metal: metals with lower boiling points generally form smaller particles (e.g., 1–50 nm for zinc), whereas metals with higher boiling points generally form larger particles (e.g. >200 nm for titanium) [21,22]. The INPMWI are not pure metals but rather metal mixtures (e.g., alloys, oxides, salts, and silicates) and carbonaceous/hydrocarbon mixtures that are either agglomerated or aggregated together in a variety of sizes and shapes [21]. Some metals that are vaporized do not condense and therefore cannot form nanometer-sized particles (e.g., mercury). 2.4. Fate of engineered nanomaterials in combustion-based municipal waste incinerators In addition to traditional waste materials, ENPs incorporated into consumer products may also contribute to INPMWI . Similar to other municipal solid waste, the effect of the incineration process on ENPs will depend on (1) the ENP chemical composition, (2) the composite into which it is incorporated (e.g., glass, plastics, paper, ceramics), (3) the incineration temperature, and (4) the duration in the incineration process. Metal ENPs exhibit lower boiling points than their bulk counterpart due to increased surface area [23]; however, combustion of metal ENPs will most likely form heteroaggregates with or without other fuel materials, which in turn will result in decreased surface area and boiling points relative to their respective bulk metals. Metal ENP stability is increased when oxidized to form metal oxides, which are heavier, less soluble, and have higher boiling points than their respective metals (e.g., aluminum versus aluminum oxide). As a result, certain metal oxide ENPs (e.g., aluminum oxide, cerium oxide) are less likely to be vaporized by the incineration process so they will likely end up in the bottom ash waste stream [24,25]. Carbon-based ENP are expected to burn completely (∼94%) so very little carbon ENP will remain in the system [26]. 2.5. Predicted INP synthesis in gasification based municipal waste incinerators In a GB-MWI system, the gasification process occurs at high temperatures and very low oxygen conditions. The highly volatile

70

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

Fig. 1. Comparison of PM10 , PM2.5 , and ultrafine particles (i.e., nanometer-sized particles) based on particle number, surface area, and mass. Reprinted from [35] with permission from HEI 2013.

Fig. 2. Removal of nanometer-sized particles from flue gas by baghouse fabric filters. Reprinted from [21] with permission from Giorgio Buonanno and Elsevier.

metals particles often aggregate (alloy) with, or vaporize and condensate on, surrounding metals and uncombusted organic material in the gasification chamber [22]. Bulk and nanometer-sized metals and metal oxides with higher boiling points usually remain unchanged. If sufficient oxygen is present, some metals (e.g., aluminum, iron, bismuth) may convert to metal oxides and become more stable. Carbon nanoparticles and organic components of nanocomposite materials have much lower boiling points, there-

fore, there is a high likelihood that these materials will convert to syngas, amorphic carbonaceous soot (though not as likely due to the substochiometric ratio of fuel:oxygen mixture), or agglomerate/aggregate to combustion by-product (bottom ash). Soot or combustion by-product may entrap nanoparticles or provide a surface to which INPMWI or condensed aerosol vapors can adhere. GB-MWI also allows for larger, uncombusted matter and larger, unmelted metals to be collected and separated as bottom ash so

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

71

it does not enter the syngas chamber and have the potential to melt, volatilize, or convert into small PM. It is possible that some of the larger, agglomerated ENPs will also separate out with the bottom ash rather than move on to the syngas burner. Smaller, lighter INPMWI may get caught in air currents and move on to the syngas burner. In integrated gasification combined cycle (IGCC) facilities, the gasification process is combined with a power generation process. After syngas is generated and cleaned, syngas is combusted in a gas turbine at high temperatures (up to 1650 ◦ C) which produces heat to generate steam that is sent to a turbine to produce electricity [19]. Most combusted fuel consists of the syngas constituents and most combustion residue consists of carbonaceous PM (e.g. fly ash), similar to gas-fired power plants [27]. It is possible that ENP and INPMWI may be transported into the syngas burner where they may alter the combustion process, such as acting as catalysts and increasing syngas combustion efficiency [28,29]. Alternatively, ENPs and INPMWI may alter PM composition, such as hetero-agglomeration (i.e., agglomerating to particles of different chemical compositions) and chemical modifications (i.e., functional group formation) [30]. Some ENPs may melt or volatilize in the syngas burner, while other ENPs may form metal oxides or micrometer-sized metal aggregates (e.g., aluminum or silver) [21,22]. As the flue gases cool down while traveling to emission control devices (Section 3), volatilized metals can condense and form new INPMWI , though the size may be large and/or agglomerated due to environmental factors.

3. Exposure 3.1. Flue gas and emissions control technologies Fine and ultrafine PM that does not partition in bottom ash is carried out with the flue gas towards the stack. Emission controls are in place to capture and reduce the release of NAAQS gases and PM into the environment as solid particulate aerosols. Increasingly efficient emission control technologies have been developed and employed to keep incinerators within compliance of increasingly stringent PM standards that protect human health (e.g., NAAQS as part of the U.S. Clean Air Act, the Danish Environmental Protection Law of 1973, EU Waste Incineration Directive (EU-WID) (2000/76/EC), and the EU Industrial Emissions (IED)(2010/75/EU)). Damgaard et al. [2] presented a life cycle assessment (LCA) of different emission control technologies that specifically highlight increased particle removal throughout the past several decades as technologies improved. In fact, current technologies are predicted to prevent the release of 99.94% of particles compared to 1973 particle emissions [2]. These results are in agreement with Arena and DiGregorio [20] where existing emissions control technologies for both CB-MWIs and GBMWIs are predicted to be highly efficient at removing numerous metals that end up in the flue gas. Emission control technologies result in high capture of chloride, cadmium, and mercury in CBMWIs, while GB-MWIs also capture chloride, potassium, sulfur, lead, and zinc with high efficiency. This results in the composition of flue gas consisting primarily of carbon (98–99% output stream, indicating high efficient energy conversion) and sulfur (2–3% output stream). All other metals constitute less than 1% of the input streams. With INPMWI being a minor contributor to PM2.5 mass values, scientists have recently started using particle number count as an improved metric for INPMWI analysis. Using this approach, Buonanno et al. [21,31] confirmed the engineering model predictions by Arena and Digregorio [20] in field measurements where emissions technologies removed >99.9% of all ultrafine particles. In fact, the use of fabric filters alone removed 99.8% of ultrafine particles

Fig. 3. Source contributions of nanometer-sized particles in the ambient atmosphere. The amount of nanometer-sized particle emitted is estimated from the survey values. (A) Nanometer-sized particle emission sources during 1996 in California’s south coast air basin that surrounds Los Angeles. (B) Nanometer-sized particle emission sources in 2005 to the European particle emissions. Reprinted from [35] with permission from David Kittelson, Win Watts, Hugo Denier van der Gon, and HEI 2013.

in flue gas (INPMWI particle concentration before filter: 2.4 × 107 particles/cm3 ; INPMWI particle concentration after filter: 3.50 × 102 particles/cm3 ) (Fig. 2) [21]. Cernuschi et al. [32] found similar results with dry process technologies, where the MWI studied released 0.5 × 104 to 1.6 × 104 particles/cm3 at the stack, which was consistently lower than ambient air concentrations (1.3–3.8 × 104 particles/cm3 ; site characterized by Lonati et al. [33]). Interestingly, lower particle number concentrations occurred at incinerators that had a lower baghouse temperature, possibly due to enhanced gasto-particle conversion in the flue gas upstream of the fabric filter and improving capture efficiency [32]. Wet emissions technologies did not appear to be as effective at reducing INPMWI releases [32,34].

72

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

Fig. 4. Relative heavy metal mass concentration (percentage of total mass) nanometer-sized particles released from a municipal waste incinerator. Reprinted from [21] with permission from Giorgio Buonanno and Elsevier.

The discrepancy between these two emissions control technologies is thought to be due to the high humidity in the wet scrubber which causes aerosolization of liquid content, hence higher particle counts since the particle counters do not distinguish between solid and liquid aerosols.

3.2. Incidental nanometer-sized particle release into the atmosphere Despite the best available emission control technologies, some INPMWI are released from MWIs into the atmosphere. These particles vary in size, shape, and composition, depending on the municipal waste feed, MWI process, and emission control technologies [21,31,32]. Once in the atmosphere, INPMWI are diluted in the ambient atmosphere. Furthermore, INPMWI also mix with other PM, ultrafine aerosols, gases, and volatile and semi-volatile organic compounds present in the ambient atmosphere. This, in turn, makes identifying the sources of specific sized PM difficult. While most pollutant release studies focus on the larger PM2.5 and PM10 , there have been limited surveys that have focused on

identifying general sources of nanometer-sized PM (termed PM1.0 [less than 1 ␮m] or PM0.1 [less than 0.1 ␮m or 100 nm]) in the atmosphere. The California South Coast Air Basin survey [35] identified sources of PM0.1 emissions in the area surrounding Los Angeles, California, USA (Fig. 3a). The calculated total PM0.1 emission in the basin was 13.25 t per day. This emissions inventory identified transportation sources (road and non-road) as the primary contributor to PM0.1 (54% of the total PM0.1 ). Combustion sources (residential, commercial, industrial, and other) contributed 23% of the PM0.1 emissions in this region. An emissions inventory by the European Commission [35] identified sources of nanometer-sized particles less than 0.3 ␮m (PM0.3 ) in Europe (Fig. 3b). Much like the California South Coast Air Basin survey, this emissions inventory identified transportation sources (i.e., on-road vehicles and other mobile sources) as the primary source of PM0.3 particles (51%). PM0.3 particles from MWIs were likely higher than PM0.3 particles from the ‘waste burning’ category (1%), but were likely much lower than PM0.3 particles from the ‘stationary fuel use’ category (32%). The caveats of these survey data are that they are regionspecific, change with time, and they have not been verified with

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

73

Fig. 5. Representative particle number concentration trends in the daily winter activities of a household. Reprinted from [38] with permission from Giorgio Buonanno and Elsevier.

field measurements [35]. Nonetheless, these surveys show the wide discrepancy between nanometer-sized particles from point sources (e.g., MWIs) and non-point sources (e.g., transportation). The actual exposure to INPMWI is difficult to measure due to the wide variety of combustion sources. In an effort to differentiate between MWI and transportation sources, Buonanno et al. [21,36] measured INPMWI released from a rural MWI in Italy. The sampling site was 200 m downwind of the MWI and 400 m downwind of a six-lane highway, making this an ideal location to differentiate INPMWI release compared to other transportation sources. The particle number concentration of INPMWI measured at the MWI stack was 2.0 × 103 particles/cm3 , consisted of complex mixtures of metals (Fig. 4), mineral compositions (e.g., oxides, salts, silicates), and shapes (square, polygons, platelets, and transparent thin plates) from 9 to 450 nm [21]. The nanometer-sized particle number concentrations from the nearby six-lane highway showed wide variations due to several factors, including time of day (e.g., morning and afternoon commuting hours), time of week (e.g., weekdays vs. weekends), weather, and seasons [21,36]. Considered a worst-exposure scenario, particle number counts during high traffic conditions were 2.0 × 105 particles/cm3 near the roadway, yet particle number counts dropped off precipitously at downwind monitoring stations so that particle number counts from the monitoring site (400 m downwind) were the same as background particle levels 400 m upwind of the highway. After monitoring the MWI for 12 months, the annual mean values at the downwind monitoring site were 8.6 × 103 ± 3.7 × 102 particles/cm3 and 31.1 ± 9.0 mg/cm3 for particle count and mass (PM10 ) concentration, respectively. These values are considered normal for rural areas. Thus, the contribution from the MWI was considered negligible compared to the highway source. Furthermore, these data demonstrate the difficulty of differentiating INPMWI once they enter the environment due to the prevalence of anthropogenic sources, especially transportation, and weather conditions. To further differentiate INPMWI and transportation sources, Buonanno and Morawska [37] compiled results from several studies to demonstrate differences in particle number concentrations in transportation microenvironments. As seen in Table 1, the INPMWI particle number concentration at the stack (5.5 × 103 particles/cm3 ) ranged between concentrations found in clean background and urban background microenvironments (3.1 × 103 and

Table 1 Median particle number concentrations of INPs in different microenvironments. Microenvironment

Median Particle Number (Particles/cm3 )

MWI Stack

5.5 × 103

Clean Background Rural Urban Background Urban

3.1 × 103 2.9 × 103 8.5 × 103 8.8 × 104

On Road Road Side Street Canyon Tunnel

4.7 × 104 3.5 × 104 4.0 × 104 9.9 × 104

Source: Buonanno and Morawska [37].

8.5 × 103 particles/cm3 , respectively). Furthermore, the application of site-specific data into air dispersion models, such as AERMOD atmospheric dispersion modeling system, will result in dilution factors that estimated ground-level particle number concentrations of INPMWI (i.e., concentrations at which workers and the public may be exposed) which will be several orders of magnitude below concentrations at the stack. In contrast, particle number concentrations of INPs in all transportation microenvironments—all of which are near ground-level—were at least an order of magnitude higher than background values. These transportation sources do not benefit from dilution factors, so the public is likely to be exposed to these INP concentrations. 3.3. Nanometer-sized particle exposure in personal microenvironments While studies have recently examined the release of nanoparticles into the atmosphere from various anthropogenic sources, few have actually examined the uptake of such particles by humans outside an occupational setting. Buonanno et al. [38] examined nanometer-sized particle exposures in microclimates for both men and women volunteers (Fig. 5). The use of personal monitors showed a wide variation of exposure throughout the day. Women who were homemakers were, on average, exposed to significantly higher particle number concentrations than men in both summer and winter (summer: 1.8 × 104 particles/cm3 (women) vs. 9.2 × 103 particles/cm3 (men); winter: 2.9 × 104 particles/cm3

74

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

(women) vs. 1.3 × 104 particles/cm3 (men)). This higher exposure was linked to cooking activities (e.g., propane-combustion particles and food aerosols) and reduced indoor ventilation, especially in winter months. Carbonaceous INPs generated from propane combustion have been shown to come in a variety of sizes and shapes [39–41]. The dose intensity (i.e., dose per unit time) for cooking activities was 8.6 and 6.6 for winter and summer, respectively. Transportation was the activity with highest exposure for men (dose intensity = 2.8). Interestingly, sleeping was also a major contributor of nanometer-sized particle exposure (approximately 20% for both men and women), due to the low concentrations over a long duration and reduced indoor ventilation. In all, time spent indoors was the major contributor to nanometer-sized particle exposures (89–97% of daily dose fraction). An additional microenvironment study by Mazaheri et al. [42] shows that children also have high indoor exposure levels, with home, school, commuting, and other sources contributed 55.3%, 35.3%, 4.5%, and 5.0%, respectively, of the children’s daily proportional uptake. While most exposure studies measure particle concentration in the vicinity of the breathing zone, few studies have measured actual inhaled concentrations of nanoparticles by humans. A study by Churg and Brauer [43] is one of the few studies that analyzed nanometer-size particle uptake into human lungs. Lung samples were from never-smokers that lived in or near Vancouver, British Columbia, Canada, a region with relatively low ambient pollutants at the time of the study (1984–1993 average: 25 ␮g PM10 and 15 ␮g PM2.5 [44]). Particle numbers in lung were surprisingly high, reaching 1 × 107 particles/g dry tissue, and nanometer-size particle number concentration generally increased with increasing airway generation, i.e., the smaller the airway the higher the particle number concentration. Interestingly, carinal mucosa (i.e., the mucosa where airways bifurcate into smaller airways) were impacted disproportionally high, with particle burdens 8–10 times higher than the immediately preceding tubular airway segment. This suggests that carinas may be a preferential target of particle toxicity. Most of the particles (>90%) examined in the airways had aerodynamic diameters less than 2.5 ␮m. The geometric mean particle diameter by lung location varied from 310 to 610 nm, with a general trend of larger particles retained in the respiratory and membranous bronchioles. Less than 15% of the particles detected were ultrafine particles (i.e., <100 nm), which were mostly spherical singlets and consisted of crystalline silica, titanium (metal and oxide forms), or iron particles (metal and oxide forms). Carbonaceous nanometer-sized particles and agglomerates/aggregates, indicators of air pollution, were not detected in Vancouver lung samples. In contrast, lung samples from highly polluted Mexico City, Mexico (the three-year average PM10 = 66 ␮g/m3 ) contained numerous carbonaceous chain aggregates and ultrafine particles, similar in composition and morphology to air samples, suggesting they were from combustion sources, such as diesel exhaust [43].

4. Toxicity of nanometer-sized particles For the general public, inhalation will be the primary route of exposure to INPMWI . Physiological studies have demonstrated that particles show differential deposition patterns when inhaled. In general, larger particles deposit in the nasopharyngeal region (upper respiratory tract), while smaller particles deposit in the tracheobronchial (middle respiratory tract) and alveolar (lower respiratory tract) regions (Fig. 6) [45]. Hence, smaller nanometer-sized particles are predicted to deposit in the bronchioles and alveoli by diffusion, however, smaller nanometer-sized particles (<20 nm) also deposit in upper airways due to dispersion [35,45]. Deposition is determined by many factors, including particle size, shape, chemical composition, agglomeration/aggregation state, and parti-

cle surface charge. Larger, agglomerated nanometer-sized particles would therefore be expected to deposit higher up in the respiratory tract. Deposition efficiency is also dependent upon several physiological factors, such as age, disease state, exercise, and oral versus nasal breathing [35]. Toxicologists and other scientists have devoted much time and attention to better understand the potential toxicological effects of nanometer-sized particles in the body. While it is understood that humans have always been exposed to natural nanoparticles, it is the ENPs and INPs that are of most concern due to the complexity of particle characteristics and unpredictable disposition properties. These complexities range from particle size, particle composition, particle dose/exposure, environmental factors, metrology, and differentiation from background PM (Table 2). To the author’s knowledge, no inhalation toxicity studies have been conducted with particulate emissions from MWIs. To establish the framework by which toxicity will be evaluated, knowledge about nanometer-sized particle uptake, toxicity, and epidemiology is needed to generate suitable risk assessments. 4.1. Toxicity of incidental nanometer-sized particles in animal models and human volunteers Most concern of late is the effects of INPs in atmospheric PM on the respiratory, cardiovascular, and nervous systems because epidemiological studies show associations between PM2.5 levels and adverse health effects in these systems [6,7]. It is thought that these health effects may be due, in part, to the nanometer-sized particle fraction within PM2.5 . In attempts to understand the role of ambient INPs on health effects, studies have been conducted where both animal models and human volunteers were exposed to concentrated ambient INPs. Below is a brief summary of available literature that focused on the effects of combustion-derived INPs. Emphasis was placed on inhalation studies (e.g., nose only or whole body chambers) with laboratory animals and on short-term exposure studies with human volunteer. The studies with human volunteers were either conducted in exposure chambers (allowing the investigators to moderate the exposure conditions) or out in the ambient atmosphere. Most peer-reviewed literature focuses on ambient INP from transportation sources and not incinerators; however, they do provide insight to better understand the potential health effects of ambient INPs. 4.1.1. Effects of nanometer-sized particles on the respiratory system Animal studies show that a 6-h inhalation exposure to labgenerated carbon nanometer-sized particles (150 ␮g/m3 ) do not produce lung inflammation; however, lung inflammation occurs when the concentration is increased 10-fold (1.7 mg/m3 ) [46]. Respiratory-compromised animals (e.g., aged, respiratory infected) show pulmonary inflammation and oxidative stress to short-term carbon nanometer-sized particle exposures [47,48]. Exposure to laboratory-generated combustion nanometer-sized particles (e.g., iron, soot) during neonatal development alters lung architecture and alveoli development in rats [49–52]. High levels (400 ␮g/m3 ; 2.0 × 105 particles/cm3 ) of larger nanometer-sized particles and fine particles (e.g., 100 nm < × < 2.5 ␮m) appear to elicit pulmonary inflammation and markers of allergic response in mice [53]. In human volunteer studies, pure carbon nanometer-sized particles do not elicit acute effects on the lungs [35], nor are effects seen with aged zinc oxide particles [54]. Concentrated ambient INP show limited effects on human volunteers [35]. Concentrated ambient INPs (1.52 × 105 particles/cm3 ) do not affect pulmonary function or inflammatory response in young healthy volunteers, but, interestingly, fine and coarse particles cause modest airway inflammation [55,56]. Bicyclists in high traffic and low traffic routes

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

75

Fig. 6. Predicted fractional deposition profile of inhaled particles in the adult human respiratory tract. Colors correspond with the region of the respiratory tract in which the particles are predicted to deposit. Reprinted from [45] with permission from Dr. Günter Oberdörster and Dr. Jack Harkema.

Table 2 Complexities related to nanometer-sized particles and resultant toxicity. NP Properties

Complexities Related with NP Properties

Particle Size

• Different sizes appear to elicit different toxicities and mechanisms of action. This means every nominal change in size can potentially have a different effect from the size above and below. • Agglomeration and aggregation of particles can also result in larger particles that may elicit different biological effects.

Particle Composition

• Toxicity study results are easily interpretable for studies using ENPs because the experimental design only accounted for the pure chemical component. In contrast, INPMWI are mixtures that vary due to several factors (e.g., incinerator conditions), and there is currently no distinction between toxicity due to individual particle components of INPMWI . The results from this sort of experimentation are difficult to interpret in the laboratory, much less the real world.

Particle Dose

• Typical high doses (i.e., unrealistic compared to actual potential exposure dose) used in laboratory studies may elicit a different toxicological effect than doses at lower, more realistic exposure concentrations. • INP exposures are likely to be chronic in nature. However, most experiments to date use acute exposure designs, requiring data to be extrapolated to represent chronic exposures.

Environmental Factors

• Temperature, wind, and humidity, among others, affect nanometer-sized particle disposition and thus exposure and effects.

Metrology

• There is no clear consensus which experimental designs and analytical equipment should be used to best characterize the nanometer-sized particle effects. Furthermore, it is still difficult to characterize nanometer-sized particle deposition in biological systems, though advances in analytical technology are being made at a rapid pace.

Differentiation from Background PM

• The prevalence of PM from a variety of natural and anthropogenic activities makes it difficult to distinguishing the specific source of each particle type. Specific experimental design, testing sites, or analytical techniques are needed to detect specific chemical signatures and differentiate the chemical entities.

(4.1 × 104 particles/cm3 and 2.7 × 104 particles/cm3 , respectively) in the Netherlands had little effect on lung function and inflammatory response [57]. In contrast, volunteers that participated in vigorous outdoor running trials near a major highway (average

particle count = 2.5 × 105 particles/cm3 ) elicit significant effects on airways, including reduced lung function, reduced alveolar nitric oxide (NO) levels, reduced nitrate levels, and increased oxidative stress markers, compared to subjects that participate in simi-

76

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

Table 3 Differences between worst case scenario (i.e., perceived lack of information) and realistic scenario used for the qualitative risk assessment. Worst Case Scenario

Realistic Scenario

MWI generate INPs of all materials at only small sizes (e.g., 1–100 nm)

INPMWI formation is driven by physicochemical characteristics of individual components (e.g., metals); INPMWI are mixtures of various sizes (e.g., tens to several hundred nm) Emissions control technologies capture >98% INPMWI Nanometer-sized particles are altered by the environment, resulting in agglomerates and modifications Nanometer-sized particles are blocked by natural biological barriers, greatly reducing entry into the body Concentrated INPs do not appear to be toxic, but high ambient (unfiltered INPs) elicit toxicological effects INPMWI are regulated under country or continent-specific directives (e.g., PM2.5, as per USEPA NAAQS guidelines)

Emissions control technologies do not capture INPMWI INPMWI are monodispersed when released into the atmosphere All INPMWI avoid natural biological barriers and enter the body Exposures to INPMWI lead to disease states There are absolutely no regulations for INPMWI anywhere

Fig. 7. Hazard assessment of nano-sized particles from MWIs: worst case scenario versus realistic case scenario. The color scheme represents the transition from low (green) to medium (yellow) to high (red) probability of exposure and/or adverse effects (for interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

lar trials near a vehicle-free area (mean particle count = 7.3 × 103 particles/cm3 ) [58,59]. These data suggest that nanometer-sized particles themselves do not appear to cause adverse effects on lung airways; however, the mixture of nanometer-sized particles and adsorbed co-pollutant(s) found in ambient air may be the sources of respiratory effects.

4.1.2. Effects of nanometer-sized particles on the cardiovascular system Acute respiratory exposures of rodents to carbon nanometersized particles cause increased heart rate and increased thrombogenic effects in the microcirculation of healthy animals [60,61]. Ambient INP (approximately 5 × 105 particles/cm3 ) from a Los Angeles freeway induce atherosclerosis more rapidly in a susceptible mouse model (ApoE knockout mice) compared to control mice and mice exposed to PM2.5 , demonstrating susceptible populations may be disproportionately affected by ambient INP [62]. Elder et al. [63] demonstrated that older rats (>21 months old), another susceptible animal model, exposed to on-road INP (1–3 × 105 particles/cm3 ) show enhanced levels of endothelin-2, which causes constriction of arteries and increases in blood pressure. Another study by Elder et al. [64] showed increased heart rate and heart

rate variability in hypertensive rats when exposed to on-road INP, suggesting an effect on the autonomic nervous system. In human volunteer studies, data suggest that there may be a link between laboratory-derived carbon nanometer-sized particles and pulmonary circulatory effects [35]. Studies with concentrated ambient INP show increases in clotting factors, vascular blood flow, and alterations in heart function [56,65]. Results from Mills et al. [65] also point to particulate components within diesel exhaust as the cause for vascular effects, though adsorbed gases cannot be ruled out. An outdoor exercise study conducted in high and low ambient particle count conditions showed an association between the high particle counts and reduced systemic vascular function and reduced reperfusion of small vessels in the forearm [58]. However, Langrish et al. [66] demonstrated that diluted diesel exhaust, wood smoke, ozone, concentrated ambient particles, engineered carbon nanoparticles, or high ambient levels of air pollution do not increase cardiac arrhythmia in healthy volunteers or patients with coronary heart disease. In all, limited research studies suggest that there may be a link between INP exposure and cardiovascular effects; however, more detailed animal and human volunteer studies need to be conducted to better define systemic versus local effects. Further studies are also needed to differentiate and identify ambient INP composition and sources (i.e., transportation vs. MWI) responsible for adverse health effects. In addition, the following studies only address acute effects. The chronic-effects of INP are still unknown. More chronic exposure animal studies may help elucidate long-term effects.

5. Human health risk assessment 5.1. Qualitative risk assessment From the data presented above, a preliminary qualitative risk assessment can be generated for INPMWI . The flow chart used for baseline risk assessments [10] can be populated with limited data provided in this manuscript. Data Collection and Evaluation: • Site Conditions: • Combustion-based incinerators combust municipal solid waste at high temperatures, high oxygen conditions. Heat is used to generate electricity. • Gasification-based incinerators heat municipal solid waste at high temperatures and low oxygen levels to convert the waste to syngas, which can then be combusted for energy or used as a feedstock for product manufacturing. • Potential Chemicals of Concern: • Mixtures of inorganic (metal) and organic (carbon) constituents.

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

• Sizes of INPMWI generated by incineration vary based on the boiling points of chemicals (lower boiling point = smaller particles; higher boiling point = larger particles), incineration conditions (temperature, oxygen), and efficiency of combustion. • Particles agglomerate or aggregate, making them larger in size (i.e., several hundred nanometers in diameter). Exposure: • Emissions control technologies remove >99.9% of INPMWI . • Particle number concentrations of INPMWI are lower at MWI stacks than particle number concentrations of other INP in the ambient atmosphere. • Particle modification (e.g., agglomeration, oxidation) and dispersion of INPMWI is driven by environmental factors (e.g., wind, temperature, humidity, season). • Particle number concentrations of INPMWI released to the ambient atmosphere are much lower than concentrations found in transportation microenvironments (i.e., at least one order of magnitude lower). • Individual exposures to nanometer-sized particles (from all sources) vary based on microenvironments. Effects • Carbon-based nanometer-sized particles cause minor respiratory effects in healthy, but respiratory-compromised animals (e.g., aged, respiratory infected) are more adversely affected. • Exposures to carbon nanometer-sized particles affect the cardiovascular system in healthy animals. Compromised animal models (atherosclerosis, hypertensive, aged) are more susceptible to cardiovascular effects of ambient INP. • Inhalation exposure to nanometer-sized particles appears not to affect human volunteers, and concentrated ambient INPs only have limited affects. Interestingly, volunteers are more affected by fine and coarse particles. Pulmonary effects are more significant in volunteers that exercised near a major highway than a vehicle-free area, suggesting the role of mixtures of nanometersized particles and adsorbed co-pollutant(s) found in ambient air. • Data suggest that there may be a link between laboratory-derived carbon nanometer-sized particles and pulmonary circulatory effects in human volunteer studies. Similarly, cardiovascular effects are seen with concentrated ambient INPs, though adsorbed gases cannot be ruled out. However, diluted diesel exhaust, wood smoke, ozone, concentrated ambient particles, engineered carbon nanoparticles, or high ambient levels of air pollution do not increase cardiac arrhythmia in healthy volunteers or patients with coronary heart disease. Qualitative risk characterization • The potential for exposure to INPMWI is low because: • Emissions control technologies are highly efficient; • Releases are lower than ambient atmospheric concentrations; • Releases are diluted several orders of magnitude before reaching ground level where exposure to the public may occur; • The majority of nanometer-sized particles in the ambient atmosphere are from transportation sources; and • Individuals are exposed to numerous nanometer-sized particles in their daily microenvironments. • The potential for adverse effects from INPMWI is moderate because:

77

• There are no toxicity data on MWI emissions and only limited pulmonary toxicity data on pure nanoparticles; • There appears to be limited toxicity due to concentrated ambient INPs, though it cannot be differentiated between solid phase nanometer-sized particles, liquid phased nanometer-sized particles (aerosols), and adsorbed gases; and • The limited data is for acute exposures. No data are available for chronic exposures to low concentrations of nanometer-sized particles. From this risk assessment flow chart, it can be conservatively concluded that INPMWI may pose a low to moderate risk to individuals. This is primarily based on the low particle number concentrations released at the MWI stacks and the large potential for dilution that may occur before reaching ground level where exposure may occur. However, lack of relevant toxicological data on nanometer-sized particle mixtures and nanometer-sized ambient PM is an uncertainty component that raises the potential risk. However, even with limited data, this conservative conclusion is a considerable improvement in defining potential risk posed by INPMWI compared to the worst case scenario posed by the perceived lack of information (Table 3 and Fig. 7). Additional studies of nanometer-sized particles from MWIs will help facilitate the transition of this qualitative risk assessment to a quantitative risk assessment. In fact, this transition is already taking place with a recent study by Scungio [67] where a quantitative lung cancer risk assessment was conducted for ultrafine particles from a waste-to-energy MWI in Italy. Using site-specific input parameters in a numerical computational fluid dynamics model, the Excess Lifetime Cancer Risk (ELCR) from ultrafine particles was below the World Health Organization’s target risk (1 × 10−5 ) [67]. 5.2. Data gaps Nanometer-sized particle toxicity is extremely complex and full of uncertainties in the laboratory, much less out in the atmosphere. The variations of particle size, shape, and composition, especially with combustion-derived INPs, makes it difficult to differentiate physical effects from chemical effects. Ambient INP chemistry can also be altered by numerous atmospheric chemistry parameters (e.g., photo-degradation, humidity, temperature, oxidation, chemical adsorption). Limited human volunteer studies have been conducted with ambient INPs, and even fewer studies have been conducted with specific nanometer-sized particles, making the use of animal models critical. There are undoubtedly other key factors that have not been identified yet that drive nanometer-sized particle fate and effects. For instance, several animal and human volunteer studies mentioned above indicated that adverse health effects may be due to adsorption of co-pollutants, not necessarily the nanoparticle or ambient INPs [58,59,65]. Predicting nanoparticle risk with current models, and not knowing these unidentified key influences, can propagate uncertainty and lead to incorrect risk estimates [45]. In addition, current epidemiological data are limited in how they can interpret the risk due to INPMWI . For example, epidemiology studies do not differentiate between nanometer-sized particles and larger fine particles when conducting analyses of PM2.5 [68,69]; they do not differentiate between individual components in the mixtures (i.e., metals vs. adsorbed gases) [6,7]; differentiating sources are difficult [7,70,71] and there is often a paucity of available information on emission from specific sources; and regional monitoring stations are used to generalize exposure at regional levels though they do not take into account the role of microenvironments on the dose- and time-response relationship of pollutants on health effects [72]. Furthermore, scientists suggest that only by examining personal daily activities can toxicologists and epidemi-

78

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79

ologists better link specific size fractions (e.g., nanometer-sized vs. micrometer-sized particles) within PM2.5 with adverse human health effects. Personal monitors, as mentioned above, can show the dose-response, activity-response, and time-response to particle exposures in microenvironments [38]. This technique can differentiate particle exposures due to specific activities, such as cooking, transportation, and work. This approach may provide the necessary resolution for epidemiological studies to identify links between sources nanometer-sized particles and health effects. In lieu of epidemiological data for nanoparticles, animal models can be used as an alternative to generate the needed information to make preliminary risk decisions. This would allow researchers to use surrogate ENPs, such as carbon nanotubes, titanium dioxide, cerium oxide, and iron oxide, to generate toxicity values and employ adaptive risk management decisions when new data are available. Unfortunately, there still is no consensus as to the correct experimental designs and analytical procedures to fully characterize nanoparticle toxicity in animal models. Oberdörster et al. [45] listed several factors that need to be addressed by the research community before good, robust data are collected for nanoparticle risk analyses. These factors include (1) metrics for accurately determining toxicity (e.g., surface area), (2) chemical composition, (3) particle size, (4) concentration (e.g., mass vs. particle count), (5) dose route (e.g., nose-only inhalation, whole body chamber, intratracheal installation), (6) exposure duration (acute vs. chronic), (7) toxicological endpoints (depends on duration), (8) volunteer/patient personal history, and (9) predisposition to disease state (e.g., age, hypertension, asthma). Without a more complete understanding of the toxicity of nanoparticles and ambient INPs, risk decisions based solely on a mass basis may be under-conservative because the wrong dose metric was used [45]. Kandlikar et al. [73] suggest using expert judgment approaches to study the degrees of consensus and disagreements between experts on the nanoparticle chemistry, fate, and toxicity. This, in turn, can provide guidance for health screening levels, while identifying data gaps and research prioritization through value of information analyses [74]. This approach fits into an adaptive risk management strategy, which can reevaluate the risk from nanometer-sized particles as more data becomes available and thus reduce risk. Acknowledgements The author would like to thank Drs. Giorgio Buonanno, Christie Sayes, and Stephen Zemba and Mr. Stephen Koo for their technical contributions and review. The author would also like to thank Ms. Lori Caranfa for creating the graphic abstract. References [1] E. Rosenthal, Europe finds clean energy in trash, but U.S. lags. The New York Times (April 12, 2010), retrieved 04/13/2014, from http://www.nytimes.com. [2] A. Damgaard, C. Riber, T. Fruergaard, T. Hulgaard, T.H. Christensen, Life-cycle-assessment of the historical development of air pollution control and energy recovery in waste incineration, Waste Manag. 20 (2010) 1244–1250. [3] P. Connett, Municipal waste incineration: a poor solution for the twenty first century, in: 4th Annual International Management Conference, Waste-To-Energy, Amsterdam, 1998. [4] Report is available at: http://www.worldbank.org/urban/solid wm/erm/ CWG%20folder/Waste%20Incineration.pdf. [5] J. Hinkle, DPH report: majority of Saugus cancer rates not unusual. Saugus Advertiser (April 6, 2016), retrieved 4/13/16; from www.saugus.wickedlocal. com. [6] U.S. Environmental Protection Agency (USEPA), Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental Protection Agency, Washington, DC EPA/600/R-08/139F, 2009, Available at http://cfpub.epa.gov/ncea/isa/recordisplay.cfm?deid=216546. [7] USEPA, Provisional Assessment of Recent Studies on Health Effects of Particulate Matter Exposure U.S. Environmental Protection Agency, Washington, DC, EPA/600/R-12/056, 2012, available at http://cfpub.epa.gov/ ncea/isa/recordisplay.cfm?deid=247132.

[8] Federal Register, National ambient air quality standards for particulate matter; Final Action. 77 Federal Register 10 (15 January 2013), pp. 3086–3287. (http://www.gpo.gov/fdsys/pkg/FR-2013-01-15/pdf/2012-30946. pdf). [9] P. Brousseau, C.J. Anderson, Nanometric aluminum in explosives, Propellants Explos. Pyrotech. 27 (5) (2002) 300–306. [10] USEPA, Risk Assessment Guidance for Superfund, Volume I, Human Health Evaluation Manual (Part A), EPA/540/1-89/002, Office of Emergency and Remedial Response, U.S. Environmental Protection Agency, Washington, D.C, 1989, available at http://www2.epa.gov/sites/production/files/2015-09/ documents/rags a.pdf. [11] USEPA, Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures. EPA/630/R-00/002, Risk Assessment Forum Technical Panel, U.S. Environmental Protection Agency, Washington DC, 2000. [12] USEPA, Framework for Cumulative Risk Assessment EPA/630/P-02/001F, Risk Assessment Forum, U.S. Environmental Protection Agency, Washington DC, 2003 https://www.epa.gov/sites/production/files/2014-11/documents/ frmwrk cum risk assmnt.pdf. [13] Report is available at: https://echa.europa.eu/documents/10162/13655/pg 15 qualitative-human health assessment documenting en.pdf. [14] R.D. Glover, J.M. Miller, J.E. Hutchison, Generation of metal nanoparticles from silver and copper objects: nanoparticle dynamics on surfaces and potential sources of nanoparticles in the environment, ACS Nano 5 (11) (2011) 8950–8957. [15] A.R. Poda, A.J. Kennedy, M.F. Cuddy, A.J. Bednar, Investigations of UV photolysis of PVP-capped silver nanoparticles in the presence and absence of dissolved organic carbon, J. Nanopart. Res. 15 (2013) 1673. [16] information is at the website: http://www.nano.gov/nanotech-101/special. [17] information is from the inventory on the website http://www. nanotechproject.org/cpi/. [18] H. Flynn, Nanotechnology Update: Corporations Up Their Spending As Revenues For Nano-enabled Products Increase, Lux Research, Boston, MA, 2013, Available at: http://www.nsf.gov/crssprgm/nano/reports/LUX14-0214 Nanotechnology%20StudyMarketResearch%20Final%2017p.pdf. [19] Gasification Technologies Council (GTC), What is Gasification?, Arlington, VA, (2015) available at http://www.gasification.org/what-is-gasification/howdoes-it-work/. [20] U. Arena, F. DiGregorio, Element partitioning in combustion- and gasification-based waste-to-energy units, Waste Manag. 33 (2013) 1142–1150. [21] G. Buonanno, L. Stabile, P. Avino, E. Belluso, Chemical, dimensional and morphometrical ultrafine particle characterization from a waste-to-energy plant, Waste Manag. 31 (11) (2011) 2253–2262. [22] S. Derrough, G. Raffin, D. Locatelli, P. Nobile, C. Durand, Behaviour of nanoparticles during high temperature treatment (incineration type), J. Phys.: Conf. Ser. 429 (2013) 012047. [23] R.A. Yetter, G.A. Risha, S.F. Son, Metal particle combustion and nanotechnology, Proc. Combust. Inst. 32 (2009) 1819–1838. [24] R.L. Lamoreaux, D.L. Hildenbrand, High-temperature vaporization behavior of oxides II. Oxides of Be, Mg, Ca, Sr, Ba, B, Al, Ga, In, Tl, Si, Ge, Sn, Pb, Zn, Cd and Hg, J. Phys. Chem. Ref. Data 16 (3) (1987) 419–443. [25] T. Walser, L.K. Limbach, R. Brogioli, E. Erismann, L. Flamigni, B. Hattendorf, M. Juchli, F. Krumeich, C. Ludwig, K. Prikopsky, M. Rossier, D. Saner, A. Sigg, S. Hellweg, D. Gunther, W.J. Stark, Persistence of engineered nanoparticles in a municipal solid-waste incineration plant, Nat. Nanotechnol. 7 (8) (2012) 520–524. [26] The report is available at: file:///C:/Users/DrJohnson/Downloads/ Nanomaterials+in+waste+incineration+and+landfills.pdf. [27] B. Hicks, S.A. McCarthy, G. Mezei, C.M. Sayes, PM1 particles at coal- and gas-fired power plant work areas, Ann. Occup. Hyg. 56 (2) (2012) 182–193. [28] Y. Gan, L. Qiao, Combustion characteristics of fuel droplets with addition of nano and micron-sized aluminum particles, Combust. Flame 158 (2) (2011) 354–368. [29] D. Ganesh, G. Gowrishankar, Effects of nano-fuel additive on emission reduction in a biodiesel fuelled CI engine, in: International Conference on Electrical and Control Engineering, Yichang, 2011. [30] E.P. Vejerano, A.L. Holder, L.C. Marr, Emissions of polycyclic aromatic hydrocarbons polychlorinated dibenzo-p-dioxins, and dibenzofurans from incineration of nanomaterials, Environ. Sci. Technol. 47 (2013) 4866–4874. [31] G. Buonanno, M. Scungio, L. Stabile, W. Tirler, Ultrafine particle emission from incinerators: the role of the fabric filter, J. Air Waste Manag. Assoc. 62 (1) (2012) 103–111. [32] S. Cernuschi, M. Guigliano, S. Ozgen, S. Consonni, Number concentration and chemical composition of ultrafine and nanoparticles from WTE (waste to energy) plants, Sci. Total Environ. 420 (2012) 319–326. [33] G. Lonati, M. Guigliano, S. Ozgen, Primary and secondary components of PM2.5 in Milan (Italy), Environ. Int. 34 (5) (2008) 665–670. [34] A. Holder, E.P. Vejerano, Z. Zhou, L.C. Marr, Nanomaterial disposal by incineration, Environ. Sci. Processes Impacts 15 (2013) 1652–1664. [35] Health Effects Institute (HEI) Review Panel on Ultrafine Particles, Understanding the health effects of ambient ultrafine particles. HEI Perspectives 3 (2013) Health Effects Institute, Boston, MA. [36] G. Buonanno, L. Stabile, P. Avino, R. Vanoli, Dimensional and chemical characterization of particles at a downwind receptor site of a waste-to-energy plant, Waste Manag. 30 (2010) 1325–1333.

D.R. Johnson / Journal of Hazardous Materials 320 (2016) 67–79 [37] G. Buonanno, L. Morawska, Ultrafine particle emission of waste incinerators and comparison to the exposure of urban citizens, Waste Manag. 37 (2015) 75–81. [38] G. Buonanno, L. Stabile, L. Morawska, Personal exposure to ultrafine particles: the influence of time-activity patterns, Sci. Total Environ. 468–469 (2014) 903–907. [39] L.E. Murr, Microstructures and nanostructures for environmental carbon nanotubes and nanoparticulate soots, Int. J. Environ. Res. Public Health 5 (5) (2008) 321–336. [40] L.E. Murr, J.J. Bang, E.V. Esquivel, P.A. Guerrero, D.A. Lopez, Carbon nanotubes nanocrystal forms, and complex nanoparticle aggregates in common fuel-gas combustion sources and the ambient air, J. Nanopart. Res. 6 (2004) 241–251. [41] L.E. Murr, K.F. Soto, K.M. Garza, P.A. Guerrero, F. Martinez, E.V. Esquivel, D.A. Ramirez, Y. Shi, J.J. Bang, J. Venzor III, Combustion-generated nanoparticulates in the El Paso, TX, USA/Juarez, Mexico metroplex: their comparative characterization and potential for adverse health effects, Int. J. Environ. Res. Public Health 3 (1) (2006) 48–66. [42] M. Mazaheri, S. Clifford, R. Jayaratne, M.A.M. Mokhtar, F. Fuoco, G. Buonanno, L. Morawska, School children’s personal exposure to ultrafine particles in the urban environment, Environ. Sci. Technol. 48 (2014) 113–120. [43] A. Churg, M. Bauer, Ambient atmospheric particles in the airways of human lungs, Ultrastruct. Pathol. 24 (6) (2000) 353–361. [44] J.R. Brook, T.F. Dann, The relationship among TSP, PM10 PM2.5, and inorganic constituents of atmospheric particulate matter at multiple Canadian locations, J. Air Waste Manag. Assoc. 47 (1997) 2–19. [45] G. Oberdörster, E. Oberdörster, J. Oberdörster, Nanotoxicology: an emerging discipline evolving from studies of ultrafine particles, Environ. Health Perspect. 113 (7) (2005) 823–839. [46] P.S. Gilmour, A. Ziesenis, E.R. Morrison, M.A. Vickers, E.M. Drost, I. Ford, E. Karg, C. Mossa, A. Schroeppel, G.A. Ferron, J. Heyder, M. Greaves, W. MacNee, K. Donaldson, Pulmonary and systematic effects of short-term inhalation exposure to ultrafine carbon black particles, Toxicol. Appl. Pharmacol. 195 (2004) 35–44. [47] A.C. Elder, R. Gelein, J.N. Finkelstein, C. Cox, G. Oberdörster, Pulmonary inflammatory response to inhaled ultrafine particles is modified by age, ozone exposure, and bacterial toxin, Inhal. Toxicol 4 (12 Suppl) (2000) 227–246. [48] A.C.P. Elder, R. Gelein, J.N. Finkelstein, C. Cox, G. Oberdörster, Endotoxin priming affects the lung response to ultrafine particles and ozone in young and old rats, Inhal. Toxicol. 12 (2000) 85–98. [49] K.E. Pinkerton, Y. Zhou, S.V. Teague, J.L. Peake, R.C. Walther, I.M. Kennedy, V.J. Leppert, A.E. Aust, Reduced lung cell proliferation following short-term exposure to ultrafine soot and iron particles in neonate rats: key to impaired lung growth? Inhal. Toxicol. 1 (2004) 73–81. [50] K.E. Pinkerton, Y. Zhou, C. Zhong, K.R. Smith, S.V. Teague, I.M. Kennedy, M.G. Menache, Mechanisms of Particulate Matter Toxicity in Neonatal and Young Adult Rat Lungs. Research Report 135, Health Effects Institute, Boston, MA, 2008. [51] C.Y. Zhong, Y.M. Zhou, K.R. Smith, I.M. Kennedy, C.Y. Cheng, A.E. Aust, K.E. Pinkerton, Oxidative injury in the lungs of neonatal rats following short-term exposure to ultrafine iron and soot particles, J. Toxicol. Environ. Health A 73 (2010) 837–847. [52] D. Lee, C. Wallis, A.S. Wexler, E.S. Schelegle, L.S. Van Winkle, C.G. Plopper, M.V. Fanucchi, B. Kumfer, I.M. Kennedy, J.K. Chan, Small particles disrupt postnatal airway development, J. Appl. Physiol. 109 (2010) 1115–1124. [53] M.T. Kleinman, A. Hamade, D. Meacher, M. Oldham, C. Sioutas, B. Chakrabarti, D. Stram, J.R. Froines, A.K. Cho, Inhalation of concentrated ambient particulate matter near a heavily trafficked road simulates antigen-induced airway responses in mice, J. Air Waste Manag. Assoc. 55 (2005) 1277–1288. [54] W.S. Beckett, D.F. Chalupa, A. Pauly-Brown, D.M. Speers, J.C. Stewart, M.W. Frampton, M.J. Utell, L.S. Huang, C. Cox, W. Zareba, G. Oberdörster, Comparing inhaled ultrafine versus fine zinc oxide particles in healthy adults: a human inhalation study, Am. J. Respir. Crit. Care Med. 171 (2005) 1129–1135. [55] J.M. Samet, D. Graff, J. Berntsen, A.J. Ghio, Y.C. Huang, R.B. Devlin, A comparison of studies on the effects of controlled exposure to fine: coarse and ultrafine ambient particulate matter from a single location, Inhal. Toxicol. 1 (19 Suppl) (2007) 29–32. [56] J.M. Samet, A. Rappold, D. Graff, W.E. Cascio, J.H. Berntsen, Y.C. Huang, M. Herbst, M. Bassett, T. Montilla, M.J. Hazucha, P.A. Bromberg, R.B. Devlin, Concentrated ambient ultrafine particle exposure induces cardiac changes in young healthy volunteers, Am. J. Respir. Crit. Care Med. 179 (2009) 1034–1042.

79

[57] M. Strak, H. Boogaard, K. Meliefste, M. Oldenwening, M. Zuurbier, B. Burnekreff, G. Hoek, Respiratory health effects of ultrafine and fine particle exposure in cyclists, Occup. Environ. Med. 67 (2010) 118–124. [58] K.W. Rundell, J.R. Hoffman, R. Caviston, R. Bulbulian, A.M. Hollenbach, Inhalation of ultrafine and fine particulate matter disrupts systemic vascular function, Inhal. Toxicol. 19 (2007) 133–140. [59] K.W. Rundell, J.B. Slee, R. Caviston, A.M. Hollenbach, Decreased lung function after inhalation of ultrafine and fine particulate matter during exercise is related to decreased total nitrate in exhaled breath condensate, Inhal. Toxicol. 20 (2008) 1–9. [60] V. Harder, P. Gilmour, B. Lentner, E. Karg, S. Takenaka, A. Ziesenis, A. Stampfl, U. Kodavanti, J. Heyder, H. Schulz, Cardiovascular responses in unrestrained WKY rats to inhaled ultrafine carbon particles, Inhal. Toxicol. 17 (2005) 29–42. [61] A. Khandoga, T. Stoeger, A.G. Khandoga, P. Bihari, E. harg, D. Ettehadieh, S. Lakatos, J. Fent, H. Schulz, F. Krombach, Platelet adhesion and fibrinogen deposition in murine microvessels upon inhalation of nanosized carbon particles, J. Thromb. Haemost. 8 (2010) 1632–1640. [62] J.A. Araujo, B. Barajas, M. Kleinman, X. Wang, B.J. Bennett, K.W. Gong, M. Navab, J. Harkema, C. Sioutas, A.J. Lusis, A.E. Nel, Ambient particulate pollutants in the ultrafine range promote early atherosclerosis and systemic oxidative stress, Circ. Res. 102 (2008) 589–596. [63] A. Elder, R. Gelein, J. Finkelstein, R. Phipps, M. Frampton, M. Utell, D.B. Kittelson, W.F. Watts, P. Hopke, C.H. Jeong, E. Kim, W. Liu, W. Zhao, L. Zhuo, R. Vincent, P. Kumarathasan, G. Oberdörster, On-road exposure to highway aerosols. 2. Exposures of aged, compromised rats, Inhal. Toxicol. 1 (16 Suppl) (2004) 41–53. [64] A. Elder, J.P. Couderc, R. Gelein, S. Eberly, C. Cox, X. Xia, W. Zareba, P. Hopke, W. Watts, D. Kittelson, M. Frampton, M. Utell, G. Oberdörster, Effects of on-road highway aerosol exposures on autonomic responses in aged, spontaneously hypertensive rats, Inhal. Toxicol. 19 (2007) 1–12. [65] N.L. Mills, M.R. Miller, A.J. Lucking, J. Beveridge, L. Flint, A.J. Boere, P.H. Fokkens, N.A. Boon, T. Sandstrom, A. Blomberg, R. Duffin, K. Donaldson, P.W. Hadoke, F.R. Cassee, D.E. Newby, Combustion-derived nanoparticluate induces the adverse vascular effects o diesel exhaust inhalation, Eur. Heart J. 32 (2011) 2660–2671. [66] J.P. Langrish, S.J. Watts, A.J. Hunter, A.S. Shah, J.A. Bosson, J. Unosson, S. Barath, M. Lundback, F.R. Cassee, K. Donaldson, T. Sandstrom, A. Blomberg, D.E. Newby, N.L. Mills, Controlled exposures to air pollutants and risk of cardiac arrhythmia, Environ. Health Perspect. 122 (7) (2014) 747–753. [67] M. Scungio, G. Buonanno, L. Stabile, G. Ficco, Lung cancer risk assessment at receptor site of a waste-to-energy plant, Waste Manag. (16) (2016) 30373–30377, http://dx.doi.org/10.1016/j.wasman.2016.07.027, pii: S0956-053X [Epub ahead of print]. [68] T. Kuwayama, C.R. Ruehl, M.J. Kleeman, Daily trends and source apportionment of ultrafine particulate mass (PM0.1) over an annual cycle in a typical California city, Environ. Sci. Technol. 47 (24) (2013) 13957–13966. [69] H.E. Wichmann, C. Spix, T. Tuch, G. Wolke, A. Peters, J. Heinrich, W.G. Keyling, J. Heyer, Daily Mortality and Fine and Ultrafine Particles in Erfurt, Germany, Part I: Roll of Particle Number and Particle Mass Research Report 98, Health Effects Institute, Cambridge, MA, 2013. [70] A.S. Kamal, A.C. Rohr, B. Mukherjee, M. Morishita, G.J. Keller, J.R. Harkema, J.G. Wagner, PM2.5-induced changes in cardiac function of hypertensive rats depend on wind direction and specific sources in Steubenville, Ohio, Inhal. Toxicol. 23 (7) (2011) 417–430. [71] A.C. Rohr, A. Kamal, M. Morishita, B. Mukherjee, G. Keller, J.R. Harkema, J.G. Wagner, Altered heart rate variability in spontaneously hypertensive rats is associated with specific particulate matter components in Detroit, Michigan, Environ. Health Perspect. 119 (4) (2011) 474–480. [72] M. Cordioli, A. Ranzi, G.A. DeLeo, P. Lauriola, A review of exposure assessment methods in epidemiological studies on incinerators, J. Environ. Public Health 2013 (2013) 129470. [73] M. Kandlikar, G. Ramachandran, A. Maynard, B. Murdock, W.A. Toscano, Health risk assessment for nanoparticles: a case for using expert judgment, J. Nanopart. Res. 9 (2007) 137–156. [74] I. Linkov, M.E. Bates, L.J. Canis, T.P. Seager, J.M. Keisler, A decision-directed approach for prioritizing research into the impact of nanomaterials on the environment and human health, Nat. Nanotechnol. 6 (12) (2011) 784–787.