-------
1
2
with Re > 100. Additional hot spots were found during expiration on the parent branch
wall downstream of the branching region.
1.0-1
2
"• 0.5-
f
"
o.o -
v
JA VP60% V
• 4
• ^f
.*•"•
:: ;-
i
.*
D t • * *
-.%v
20 40 60 80 100 120 140 160 180 200
Penetration Volume (ml_)
Source: Adapted with permission of Health Effects Institute (Ultman et al.. 2004)
Figure 5-5 Ozone uptake fraction as a function of volumetric penetration (Vp)
in a representative subject. Each point represents the O3 uptake of
a bolus inspired through a mouthpiece by the subject. The
volumes, VUA and VD, are the volume of the upper airways and
anatomical dead space, respectively, and VP50% is the Vp at which
50% of the inspired bolus was absorbed. In 47 healthy subjects,
Ultman et al. (2004) found that VP50% was well correlated with VD
and better correlated with the volume of the conducting airways,
i.e., VD minus VUA.
4
5
6
1
Overall O3 inhalation uptake in humans is over 80% efficient, but the exact efficiency
that determines how much O3 is available at longitudinally distributed compartments in
the lung is sensitive to changes in VT, fB, and to a minor extent, exposure time.
Decreased fB at a fixed penetration volume will shift the O3 uptake from the upper
airways to the central airways and respiratory airspaces.
9
10
5.2.2.5 Mode of Breathing
Ozone uptake and distribution is sensitive to the mode of breathing. Variability in TB
airways volume had a weaker influence on O3 absorption during nasal breathing
compared to oral breathing. This could be a result of O3 scrubbing in the nasal
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1 passageways that are bypassed by oral breathing. Studies by Ultman and colleagues using
2 bolus inhalation demonstrated that O3 uptake fraction was greater during nasal breathing
3 than during oral breathing at each Vp (e.g. 0.90 during nasal breathing and 0.80 during
4 oral breathing at 150 mL/s and 0.45 during nasal breathing and 0.25 during oral breathing
5 at 1,000 mL/s) (Nodelman and Ultman. 1999; Kabeletal.. 1994; Ultman et al.. 1994).
6 Therefore, oral breathing results in deeper penetration of O3 into the RT with a higher
7 absorbed fraction in the URT, TB, and alveolar airways (NodeIman and Ultman. 1999).
8 Similar results were obtained from O3 uptake studies in dogs (Yokoyama and Frank.
9 1972). Earlier human studies suggesedt that oral or oronasal breathing results in a higher
10 O3 uptake efficiency than nasal breathing ("Wiester et al.. 1996c: Gerrity et al.. 1988):
11 however the difference observed between inspired O3 taken up during oral versus nasal
12 breathing may not be biologically significant. These human studies measured total RT
13 absorption after continuous O3 exposure using a pharyngeal sampling tube, which may
14 decrease sensitivity and lead to measurement errors. Overall, the mode of breathing may
15 have little effect of the RT uptake efficiency, but does play an important role in the
16 distribution of O3 deposited in the distal airways.
5.2.2.6 Interindividual Variability in Dose
17 Similarly exposed individuals vary in the amount of actual dose delivered to the LRT
18 (Santiago etal., 2001; Rigas et al., 2000; BushetaL 1996). Interindividual variability
19 accounted for between 10-50% of the absolute variability in O3 uptake measurements
20 (Santiago et al.. 2001; Rigas et al.. 2000). When concentration, time, and MV were held
21 constant, fractional absorption ranged from 0.80 to 0.91 (Rigas et al.. 2000). It has been
22 hypothesized that interindividual variation in O3 induced response such as FEVi is the
23 result of interindividual variation in delivered dose or regional O3 uptake among exposed
24 individuals.
25 Recent studies have reiterated the importance of intersubject variation in O3 uptake. The
26 intersubject variability in nasal O3 uptake determined by Sawyer et al. (2007) ranged
27 from 26.8 to 65.4% (pre- and post-exercise). A second study investigating the use of the
28 CO2 expirogram to quantify pulmonary responses to O3 found that intersubject
29 variability accounted for 50% of the overall variance in the study (Taylor et al.. 2006).
30 Variability in local dose may be attributed to differences in the pulmonary physiology,
31 anatomy, and biochemistry. Since the TB airways remove the majority of inhaled O3
32 before it reaches the gas exchange region, the volume and surface area of the upper
33 airways will influence O3 uptake. Models predicted that fractional O3 uptake and PAR
34 dose (flux of O3 to the PAR surfaces divided by exposure concentration) increase with
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1 decreasing TB volume and decreasing TB region expansion. On the contrary, alveolar
2 expansion had minimal effect on uptake efficiency as relatively little O3 reaches the
3 peripheral lung (Bushet al.. 2001; Overton et al., 1996). Ozone uptake was virtually
4 complete by the time O3 reaches the alveolar spaces of the lung (Postlethwait et al..
5 1994). Experimental studies have found that differences in TB volumes may account for
6 75% of the variation in absorption between subjects (Hitman et al.. 2004). In support of
7 this concept, regression analysis showed that O3 absorption was positively correlated
8 with anatomical dead space (VD) and TB volume (i.e., VD minus VURT), but not total lung
9 capacity (TLC), forced vital capacity (FVC), or functional residual capacity (FRC)
10 (Ulttnan et al.. 2004: Bushetal.. 1996: Huetal.. 1994: Postlethwait et al.. 1994).
11 Variability in VD was correlated more with the variability in the TB volume than the URT
12 volume. Similarly, uptake was correlated with changes in individual bronchial cross-
13 sectional area, indicating that changes in cross-sectional area available for gas diffusion
14 are related to overall O3 retention (Reeser et al.. 2005: Ultman et al.. 2004). These studies
15 provide support to the pulmonary physiology, especially the TB volume and surface area,
16 playing a key role in variability of O3 uptake between individuals.
17 When absorption data were normalized to Vp/VD, variability attributed to gender
18 differences were not distinguishable (Bush etal.. 1996). However, variability due to age
19 has been predicted. Overton and Graham (1989) predicted that the total quantity of O3
20 absorbed per minute increased with age from birth to adulthood. This model predicted
21 that the LRT distribution of absorbed O3 and the CAR O3 tissue dose were not sensitive
22 to age during quiet breathing. However, during heavy exercise or work O3 uptake was
23 dependent on age. A physiologically based pharmacokinetic model simulating O3 uptake
24 predicted that regional extraction of O3 was relatively insensitive to age, but extraction
25 per unit surface area was two- to eightfold higher in infants compared to adults, due to
26 the fact that children under age 5 have much a much smaller airway surface area in the
27 extrathoracic (nasal) and alveolar regions (Sarangapani et al.. 2003).
28 Smoking history, with its known increase in mucus production, was not found to
29 significantly affect the fractional uptake of a bolus dose of O3 in apparently healthy
30 smokers with limited smoking history (Bates etal.. 2009). Despite similar internal O3
31 dose distribution, the smokers exhibited greater pulmonary responses to O3 bolus
32 exposures, measured as FEVi decrements and increases in the normalized slope of the
33 alveolar plateau (SN). This was contrary to previous studies conducted in smokers with a
34 greater smoking history that found decreased O3 induced decrements in FEVi in smokers
35 during continuous O3 exposure (Frampton et al.. 1997b: Emmons and Foster. 1991).
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5.2.2.7 Physical Activity
1 Exercise increases the overall exposure of the lung to inhaled contaminants due, in most
2 part, to the increased intake of air. As exercise increases from a low to moderate level, VT
3 increases. This increase in VT is achieved by encroaching upon both the inspiratory and
4 expiratory reserve volumes of the lung (Dempsey et al., 1990). After VT reaches about
5 50% of the vital capacity, generally during heavy exercise, further increases in ventilation
6 are achieved by increasing fB. Ventilatory demands of heavy exercise require airway flow
7 rates that often exceed 10 times resting levels and VT that approach 5 times resting levels
8 (Dempsev et al.. 2008).
9 This increase in VT and flow associated with exercise in humans shifts the O3 dose
10 further into the periphery of the RT causing a disproportionate increase in distal lung
11 dose. In addition to increasing the bulk transport of O3 into the lung, exercise also leads
12 to a switch from nasal to oronasal breathing. Higher ventilatory demand necessitates a
13 lower-resistance path through the mouth. Modeling heavy exercise by increasing
14 ventilatory parameters from normal respiration levels predicted a 10-fold increase in total
15 mass uptake of O3 (Miller etal.. 1985). This model also predicted that as exercise and
16 ventilatory demand increased the maximum tissue dose moved distally into the RT
17 (Figure 5-6). By increasing flow to what is common in moderate exercise (respiratory
18 flow = 750 -1,000 mL/s compared to 250 mL/s at rest), the URT absorbed a smaller
19 fraction of the O3 (-0.50 at rest to 0.10 at exercise); however, the trachea and more distal
20 TB airways received higher doses during exercise than rest (0.65 absorbed in the lower
21 TB airways, and 0.25 absorbed in the alveolar zone with exercise compared to 0.5 in the
22 TB with almost no O3 reaching the alveolar zone at rest) (Hu etal.. 1994). The same shift
23 in the O3 dose distribution more distally in the lung occurred in other studies mimicking
24 the effects of exercise (Nodehnan and Ultman, 1999). Also, LRT uptake efficiency was
25 sensitive to age only under exercise conditions (Overton and Graham. 1989). The total
26 quantity of O3 absorbed per minute was predicted to increase with age during heavy work
27 or exercise. A recent study by Sawyer et al. (2007) showed that doubling minute
28 ventilation led to only a 1.6-fold higher delivered dose rate of O3 to the lung. Past models
29 have predicted the increase in uptake during exercise is distributed unevenly in the RT
30 compartments and regions. Tissue and mucus layer dose in the TB region increased ~1.4-
31 fold during heavy exercise compared to resting conditions, whereas the alveolar region
32 surfactant and tissue uptake increased by factors of 5.2 and 13.6, respectively (Miller et
33 al.. 1985).
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4 8 12 16 20
AIRWAY GENERATION |Z)
-TB-
Source: Reprinted with permission. (Miller et al.. 1985)
Figure 5-6 Modeled effect of exercise on tissue dose of the LRT. Curve 1: VT =
500 ml_; fB = 15 breaths/min. Curve 2: VT = 1,000 ml_; fB = 15
breaths/min. Curve 3: VT = 1,750 ml_; fB = 20.3 breaths/min. Curve 4:
VT = 2,250 ml_; fB = 30 breaths/min. TB = tracheobronchial region; P
= pulmonary region.
1
2
3
4
5
6
7
8
9
10
11
12
5.2.2.8 Summary
In summary, O3 uptake is affected by complex interactions between a number of factors
including RT morphology, breathing route, frequency, and volume, physicochemical
properties of the gas, physical processes of gas transport, as well as the physical and
chemical properties of the ELF and tissue layers. The role of these processes varies
throughout the length of the RT and as O3 moves from the gas into liquid compartments
of the RT. The primary uptake site of O3 delivery to the lung epithelium is believed to be
the CAR, however inhomogeneity in the RT structure may affect the dose delivered to
this target site with larger path lengths leading to smaller locally delivered doses. Recent
studies have provided evidence for hot spots of O3 flux around bifurcations in airways.
Experimental studies and models have suggested that the net O3 dose gradually decreases
distally from the trachea toward the end of the TB region and then rapidly decreases in
the alveolar region. However, the tissue O3 dose is low in the trachea, increases to a
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1 maximum in the terminal bronchioles and the CAR, and then rapidly decreases distally
2 into the alveolar region.
3 O3 uptake efficiency is sensitive to a number of factors. Fractional absorption will
4 decrease with increased flow and increase proportional to VT, so that at a fixed MV,
5 increasing VT (or decreasing fB) drives O3 deeper into the lungs and increases total
6 respiratory uptake efficiency. Individual total airway O3 uptake efficiency is also
7 sensitive to large changes in O3 concentration, exposure time, and MV. Major sources of
8 variability in absorption of O3 include O3 concentration, exposure time, fB, MV, and VT,
9 but the interindividual variation is the greatest source of variability uptake efficiency. The
10 majority of this interindividual variability is due to differences in TB volume and surface
11 area.
12 An increase in VT and fB are both associated with increased physical activity. These
13 changes and a switch to oronasal breathing during exercise results in deeper penetration
14 of O3 into the lung with a higher absorbed fraction in the ET, TB, and alveolar airways.
15 For these reasons, increased physical activity acts to move the maximum tissue dose of
16 O3 distally into the RT and into the alveolar region.
5.2.3 Ozone Reactions and Reaction Products
17 Ozone dose can be examined by the chemical reactions or the products of these reactions
18 that result from O3 exposure. Since O3 is chemically reactive with a wide spectrum of
19 biomolecules, it is not feasible to delineate its many reaction products. Measurements of
20 reaction product formation have included either the loss of a specific molecule and
21 appearance of plausible products, or the addition of O3 -derived oxygen to biomolecules
22 through the use of oxygen-18 labeling. In vitro exposure of ELF showed that O3
23 disappearance from the gas phase depends on the characteristics of the ELF substrates
24 (Postlethwait et al.. 1998: Huetal.. 1994).
25 For O3 to gain access to the underlying cellular compartments, O3 must dissolve at the
26 air-liquid interface of the airway surface and travel through the ELF layer. The ELF is
27 comprised of the airway surface lining that includes the periciliary sol layer and
28 overlying mucus gel layer, and the alveolar surface lining that includes the subphase of
29 liquid and vesicular surfactant and the continuous surfactant monolayer (Bastacky et al..
30 1995). There is a progressive decrease in ELF thickness and increase in interfacial
31 surface with progression from the large airways to the alveolus (Mercer et al.. 1992).
32 Some cells, such as macrophages, may protrude into the gas phase, allowing for direct
33 contact between O3 and cell membranes. The progressive thinning of the ELF while
34 moving further down the RT decreases the radial distance O3 must travel to reach the
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1 cellular tissue layer. A computational fluid dynamics model was able to predict
2 experimentally measured O3 uptake, but only with nasal mucus layer thickness
3 considered (Cohen-Hubal et al.. 1996). reaffirming the importance of the resistance
4 imparted by the ELF layer in dose and lesion patterns in the nasal passage.
5 Taking into account the high reactivity and low water solubility of O3, calculations
6 suggest that O3 will not penetrate ELF layers greater than 0.1 (im without being
7 transformed to other more long-lived reactive species, thus initiating a reaction cascade
8 (Pryor. 1992). However, the surfactant layer in the pulmonary region becomes ultrathin,
9 possibly allowing for direct interaction of O3 with the underlying epithelial cells. One
10 study measured pulmonary liquid lining thickness over relatively flat portions of the
11 alveolar wall to be 0.14 (im, to be 0.89 (im at the alveolar wall junctions, and 0.09 (im
12 over the protruding features (Bastacky et al.. 1995). Still, the ELF should be considered
13 an important target for O3 and the resulting secondary oxidation products should be
14 considered key mediators of toxicity in the airways (role of reaction products in O3
15 induced toxicity is discussed in Section 5.3). Model calculations of the nasal cavity based
16 on diffusion equations and reaction rates of O3 with model substrates predict an O3
17 penetration distance (0.5 (im) less than the thickness of the mucus layer (10 (im)
18 (Santiago et al.. 2001). Experimental support for this concept comes from several studies
19 which measured the total oxygen-addition product of O3 reactions in the airways through
20 the use of oxygen-18 labeled O3. Fiigh concentrations of O3 reaction products were found
21 in the bronchoalveolar lavage (BAL) mucus and surfactant providing evidence that O3
22 reacts at the air-liquid interface. Thus, O3 may cause injury by direct reaction with
23 constituents of the lining layer, with cells protruding from it and in some cases with cells
24 underlying the lining fluid. The reaction cascade resulting from the interaction of O3 with
25 ELF substrates acts to carry the oxidative burden deeper into the tissues.
26 Ozone may interact with many of the components in the ELF including phospholipids,
27 neutral lipids, free fatty acids, proteins, and low molecular weight antioxidants (Perez-
28 Gil 2008; Uppuetal.. 1995). It was estimated that 88% of the O3 that does not come in
29 contact with antioxidants will react with unsaturated fatty acids in the ELF including
30 those contained within phospholipids or neutral lipids (Uppu et al.. 1995). Ozone reacts
31 with the double bond of lipids such as unsaturated fatty acids, a large component of ELF,
32 to form stable and less reactive ozonide, aldehyde, and hydroperoxide reaction products
33 via chemical reactions such as the Criegee ozonolysis mechanism (Figure 5-7) (Pryor et
34 al., 1991). Lipid ozonation products, such as the aldehydes hexanal, heptanal, and
35 nonanal, have been recovered after O3 exposure in human BAL fluid (BALF), rat BALF,
36 isolated rat lung, and in vitro systems (Frampton et al., 1999; Postlethwait et al., 1998;
37 Pryor et al.. 1996). Nonanal has been suggested as a relatively specific biomarker for O3
38 exposure since the monounsaturated fatty acid parent compound, oleic acid, does not
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1
2
3
4
5
6
7
8
9
10
11
undergo autoxidation (Pryor et al.. 1996). Adducts of the aldehyde 4-hydroxynonenal
were found in human alveolar macrophages after O3 exposure (Hamilton et al.. 1998).
Polyunsaturated fatty acid (PUFA) reactions are limited by the availability of O3 since
lipids are so abundant in the ELF. Yields of O3-induced aldehydes were increased by the
decrease in other substrates such as ascorbic acid (AH2) (Postlethwait et al., 1998). Free
radicals are also generated during O3-mediated oxidation reactions with PUFA (Pryor.
1994). These reactions are reduced by the presence of the lipid-soluble free radical
scavenger a-tocopherol (a-TOH) (Prvor. 1994: Fujitaetal.. 1987: Prvor. 1976). PUFA
reactions may not generate sufficient bioactive materials to account for acute cell injury,
however only modest amounts of products may be necessary to induce cytotoxicity
(Postlethwait and Ultman. 2001: Postlethwait et al.. 1998).
A
RHC =
PUFA
either in
the >
absence
ofH2O
CH +
RHC'
\
03 »
ozone
CH—
1
- RHC —
1
CH— > RHC = O — O + RHC = O
trioxolane carbonyl oxide
or in the
presence
ofH2O
Criegee ozonide
/OH
> RHC > RHC = (
XOOH aldehyde
hydroxyhydroperoxy cpd.
aldehyde
3 + H2O2
hydrogen
peroxide
Source: U.S. EPA (2QQ6b)
Figure 5-7 Schematic overview of ozone interaction with PUFA in ELF and
lung cells. It should be noted that not all secondary reaction
products are shown.
12
13
14
15
16
17
18
19
20
21
Cholesterol is the most abundant neutral lipid in human ELF. Reaction of cholesterol with
O3 results in biologically active cholesterol products such as the oxysterols, (3-epoxide
and 6-oxo-3,5-diol (Murphy and Johnson. 2008: Pulfer etal.. 2005: Pulfer and Murphy.
2004). Product yields will depend on ozonolysis conditions, however cholesterol
ozonolysis products were formed in similar abundance to phospholipid-derived
ozonolysis products in rat ELF (Pulfer and Murphy. 2004).
The ELF also contains proteins present in blood plasma as well as proteins secreted by
surface epithelial cells. Ozone reactions with proteins have been studied by their in vitro
reactions as well as reactions of their constituent amino acids (the most reactive of which
are cysteine, histidine, methionine, tyrosine, and tryptophan). Ozone has been shown to
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September 2011
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1 preferentially react with biomolecules in the following order: thiosulfate > ascorbate >
2 cysteine ~ methionine > glutathione (Kanofsky and Sima. 1995). Rate constants for the
3 reaction of amino acids with O3 vary between investigations due to differing reaction
4 conditions and assumptions; however aliphatic amino acids consistently were very slow
5 to react with O3 (e.g., alanine: 25-100 moles/L/sec) (Kanofsky and Sima. 1995;
6 Ignatenko and Cherenkevich. 1985; Pryor et al.. 1984; Hoigne and Bader. 1983). Uppu et
7 al. (1995) predicted that 12% of inhaled O3 that does not react with antioxidants will
8 react with proteins in the ELF, whereas 88% will react with PUFAs.
9 Reactions of ozone with low molecular weight antioxidants have been extensively
10 studied. The consumption of antioxidants such as uric acid (UA), ascorbate (AH2), and
11 reduced glutathione (GSH) by O3 was linear with time and positively correlated with
12 initial substrate concentration and chamber O3 concentration (Mudway and Kelly. 1998;
13 Mudway et al.. 1996). Endogenous antioxidants are present in relatively high
14 concentrations in the ELF of the human TB airways and display high intrinsic reactivities
15 toward O3, but do not possess equal O3 reactivity. In individual and in limited composite
16 mixtures, UA was the most reactive antioxidant tested, followed by AH2 (Mudway and
17 Kelly. 1998). GSH was consistently less reactive than UA or AH2 (Mudway and Kelly.
18 1998; Mudway et al.. 1996; Kanofsky and Sima. 1995). To quantify these reactions,
19 Kermani et al. (2006) recently evaluated the interfacial exposure of aqueous solutions of
20 UA, AH2, and GSH (50-200 (iM) with O3 (1-5 ppm). Similar to the results of Mudway
21 and Kelly (1998). this study found the hierarchy in reactivity between O3 and these
22 antioxidants to be UA>AH2»GSH. UA and AH2 shared a 1:1 stoichiometry with O3,
23 whereas 2.5 moles of GSH were consumed per mole of O3. Using these stoichiometries,
24 reaction rate constants were derived (S.SxlO4]^"1 sec"1, S.Sx^M"1 sec"1, and 57.5 M"
25 ° 75/sec [20.9 M"1 sec"1] for the reaction of O3 with UA, AH2, and GSH, respectively).
26 These values are similar to those derived from data presented in Mudway and Kelly
27 (1998). Other studies reported reactive rate constants that are two to three orders of
28 magnitude larger, however these studies used higher concentrations of O3 and
29 antioxidants under less physiologically relevant experimental conditions (Kanofsky and
30 Sima. 1995: Giamalva et al.. 1985: Pryor etal.. 1984).
31 A series of studies used new techniques to investigate the reaction products resulting
32 from initial air-liquid interface interactions of O3 with ELF components (e.g.,
33 antioxidants and proteins) in ~1 millisecond (Enami et al.. 2009a. b, c, 2008a. b).
34 Solutions of aqueous UA, AH2, GSH, a-TOH, and protein cysteines (CyS) were sprayed
35 as microdroplets in O3/N2 mixtures at atmospheric pressure and analyzed by electrospray
36 mass spectrometry. These recent studies demonstrated different reactivity toward AH2,
37 UA, and GSH by O3 when the large surface to volume ratio of microdroplets promote an
38 interfacial reaction compared to previous studies using bulk liquid phase bioreactors, thus
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1 supporting the relevance of reactions between gas phase O3 and antioxidants found in the
2 ELF.
3 As was seen in previous studies (Kermani et al.. 2006; Kanofsky and Sima. 1995). the
4 hierarchy of reactivity of these ELF components with O3 was determined to be AH2 ~
5 UA > CyS > GSH. There was some variance between the reaction rates and product
6 formation of UA, AH2, and GSH with O3 as investigated by Enami et al. versus O3
7 reacting with bulk liquid phase bioreactors as described previously. UA was more
8 reactive than AH2 toward O3 in previous studies, but in reactions with O3 with
9 microdroplets, these antioxidants had equivalent reactivity (Enami et al.. 2008b). As O3 is
10 a kinetically slow one-electron acceptor but very reactive O-atom donor, products of the
11 interaction of O3 with UA, AH2, GSH, CyS, and a-TOH result from addition of n O-
12 atoms (« = 1-4). These products included epoxides (e.g., U-O"), peroxides (e.g. U-O2"),
13 and ozonides (e.g., U-O3"). For instance, GSH was oxidized to sulfonates (GSO3 YGSO32
14 ), not glutathione disulfide (GSSG) by O3 (Enami et al., 2009b). However, it is possible
15 that other oxidative species are oxidizing GSH in vivo, since sulfonates are not detected
16 in O3 exposed ELF whereas GSSG is. This is also supported by the fact that O3 is much
17 less reactive with GSH than other antioxidants, such that < 3% of O3 will be scavenged
18 by GSH when in equimolar amounts with AH2 (Enami et al.. 2009b).
19 Ozonolysis product yields and formation were affected by pH. Acidified conditions (pH ~
20 3-4), such as those that may result from acidic particulate exposure or pathological
21 conditions like asthma (pH ~ 6), decreased the scavenging ability of UA and GSH for O3;
22 such that at low pH, the scavenging of O3 must be taken over by other antioxidants, such
23 as AH2 (Enami et al.. 2009b. 2008b). Also, under acidic conditions (pH ~ 5), the
24 ozonolysis products of AH2 shifted from the innocuous dehydroascorbic acid to the more
25 persistent products, AH2 ozonide and threonic acid (Enami et al.. 2008a). It is possible
26 that the acidification of the ELF by acidic copollutant exposure will increase the toxicity
27 of O3 by preventing some antioxidant reactions and shifting the reaction products to more
28 persistent compounds.
29 In a red blood cell (RBC) based system, AH2 augmented the in vitro uptake of O3 by six
30 fold, as computed by the mass balance across the exposure chamber (Ballinger et al..
31 2005). However, estimated in vitro O3 uptake was not proportional to the production of
32 O3-derived aldehydes from exposing O3 to RBC membranes (Ballinger et al.. 2005). In
33 addition, O3 induced cell membrane oxidation which required interactions with AH2 and
34 GSH, but not UA or the vitamin E analog Trolox. Further, aqueous phase reactions
35 between O3 and bovine serum albumin did not result in membrane oxidation (Ballinger et
36 al.. 2005). The presence of UA or bovine serum albumin protected against lipid and
37 protein oxidation resulting from the reaction of O3 and AH2 (Ballinger et al.. 2005). This
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1 study provided evidence that antioxidants may paradoxically facilitate O3-mediated
2 damage. This apparent contradiction should be viewed in terms of the concentration-
3 dependent role of the ELF antioxidants. Reactions between O3 and antioxidant species
4 exhibited a biphasic concentration response, with oxidation of protein and lipid occurring
5 at lower, but not higher, concentrations of antioxidant. In this way, endogenous reactants
6 led to the formation of secondary oxidation products which were injurious and also led to
7 quenching reactions which were protective. Moreover, the formation of secondary
8 oxidation products mediated by some antioxidants was opposed by quenching reactions
9 involving other antioxidants.
10 Alterations in ELF composition can result in alterations in O3 uptake. Bolus O3 uptake in
11 human subjects can be decreased by previous continuous O3 exposure (120-360 ppb),
12 possibly due to depletion of compounds able to react with O3 (Rigas etal.. 1997;
13 Asplund et al.. 1996). Conversely, O3 (360 ppb) bolus uptake was increased with prior
14 NO2 (360-720 ppb) or SO2 (360 ppb) exposure (Rigas etal.. 1997). It was hypothesized
15 that this increased fractional absorption of O3 could be due to increased production of
16 reactive substrates in the ELF due to oxidant-induced airway inflammation.
17 Besides AH2, GSH and UA, the ELF contains numerous antioxidant substances that
18 appear to be an important cellular defense against O3 including a-TOH, albumin,
19 ceruloplasmin, lactoferrin, mucins, and transferrin (Mudway et al.. 2006; Freed et al..
20 1999). The level and type of antioxidant present in ELF varies between species, regions
21 of the RT, and can be altered by O3 exposure. Mechanisms underlying the regional
22 variability are not well-understood. It is thought that both plasma ultrafiltrate and locally
23 secreted substances contribute to the antioxidant content of the ELF (Mudway et al..
24 2006; Freed etal.. 1999). In the case of UA, the major source appears to be the plasma
25 (Peden etal.. 1995). Repletion of UA in nasal lavage fluid was demonstrated during
26 sequential nasal lavage in human subjects (Mudway et al.. 1999a). When these subjects
27 were exposed to 200 ppb O3 for 2 hours while exercising, nasal lavage fluid UA was
28 significantly decreased while plasma UA levels were significantly increased (Mudway et
29 al.. 1999a). The finding that UA, but not AH2 or GSH, was depleted in nasal lavage fluid
30 indicated that UA was the predominant antioxidant with respect to O3 reactivity in the
31 nasal cavity (Mudway et al.. 1999a). In addition, concentrations of UA were increased by
32 cholinergic stimulation of the airways in exercising human subjects exposed to 400 ppb
33 O3 for 2 hours, which suggested that increased mucosal gland secretions were an
34 important source (Peden etal.. 1995). Using the O3-specific antioxidant capacity assay on
35 human nasal lavage samples, Rutkowski et al. (2011) concluded that about 30% of the
36 antioxidant capacity of the nasal liquid lining layer was attributed to UA activity. This
37 assay predicted that more than 50% of the subject-to-subject differences in antioxidant
38 capacity were driven by differences in UA concentration. However, day-to-day within-
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1 subject variations in measured antioxidant capacity were not related to the corresponding
2 variations in UA concentration in the nasal lavage fluid. Efforts to identify the
3 predominant antioxidant(s) in other RT regions besides the nasal cavity have failed to
4 yield definitive results. However, in human BALF samples, the mean consumption of
5 AH2 was greater than UA (Mudway et al.. 1996).
6 Regulation of AH2, GSH and a-TOH concentrations within the ELF is less clear than that
7 of UA (Mudway et al.. 2006). In a sequential nasal lavage study in humans, wash-out of
8 AH2 and GSH occurred, indicating the absence of rapidly acting repletion mechanisms
9 (Mudway et al.. 1999a). Other studies demonstrated increases in BALF GSH and
10 decreases in BALF and plasma AH2 levels several hours following O3 exposure (200 ppb
11 for 2 h, while exercising) (Mudway et al.. 2001; Blomberg et al.. 1999; Mudway et al..
12 1999b). Furthermore, high levels of dehydroascorbate, the oxidized form of AH2, have
13 been reported in human ELF (Mudway et al.. 2006). Other investigators have
14 demonstrated cellular uptake of oxidized AH2 by several cell types leading to
15 intracellular reduction and export of reduced AH2 (Welch etal.. 1995). Studies with rats
16 exposed to 0.4-1.1 ppm O3 for 1-6 hours have shown consumption of AH2 that correlates
17 with O3 exposure (Gunnison and Hatch. 1999; Gunnison et al.. 1996; Vincent et al..
18 1996a).
19 ELF exists as a complex mixture, thus it is important to look at O3 reactivity in substrate
20 mixtures. Individual antioxidant consumption rates decreased as the substrate mixture
21 complexity increased (e.g., antioxidant mixtures and albumin addition) (Mudway and
22 Kelly. 1998). However, O3 reactions with AH2 predominated over the reaction with
23 lipids, when exposed to substrate solution mixtures (Postlethwait et al.. 1998). It was
24 suggested that O3 may react with other substrates once AH2 concentrations within the
25 reaction plane fall sufficiently. Additionally, once AH2 was consumed, the absorption
26 efficiency diminished, allowing inhaled O3 to be distributed to more distal airways
27 (Postlethwait et al.. 1998). Multiple studies have concluded O3 is more reactive with AH2
28 and UA than with the weakly reacting GSH (or cysteine or methionine) or with amino
29 acid residues and protein thiols (Kanofsky and Sima. 1995; Cross et al.. 1992).
30 In addition to reactions with components of the ELF, O3 may react with plasma
31 membranes of cells which reside in the RT. Eicosanoids are an important class of
32 secondary oxidation products which may be formed rapidly by this mechanism.
33 Eicosanoids are metabolites of arachidonic acid, a 20-carbon PUFA, which is released
34 from membrane phospholipids by phospholipase A2-mediated catalysis. Activation of
35 phospholipase A2 occurs by several cell signaling pathways and may be triggered by O3-
3 6 mediated lipid peroxidation of cellular membranes (Rashba-Step et al.. 1997).
37 Additionally, cellular phospholipases A2, C and D may be activated by lipid ozonation
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1 products (Kafoury et al., 1998). While the conversion of arachidonic acid to
2 prostaglandins, leukotrienes and other eicosanoid products is generally catalyzed by
3 cyclooxygenases and lipoxygenases, non-enzymatic reactions also occur during oxidative
4 stress leading to the generation of a wide variety of eicosanoids and reactive oxygen
5 species. Further, the release of arachidonic acid from phospholipids is accompanied by
6 the formation of lysophospholipids which are precursors for platelet activating factors
7 (PAFs). Thus, formation of eicosanoids, reactive oxygen species and PAFs accompanies
8 Os-mediated lipid peroxidation.
5.2.3.1 Summary
9 The ELF is a complex mixture of lipids, proteins, and antioxidants that serve as the first
10 barrier and target for inhaled O3 (Figure 5-8). The thickness of the lining fluid and mucus
11 layer is an important determinant of the dose of O3 to the tissues. The antioxidant
12 substances present in the ELF appear in most cases to limit interaction of O3 with
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Ozone
Mucus Layer
Mudns
Phospholipids
Liquid Layer Antioxidants
Uric Acid, Ascofbate,
Giutathione, a-To-copherol
ELF Macromolecules
Surfactant components
e.g. proteins, phospholipids/
cholesterol, CCSP,
Albumin, Hyaluronan, SP-A
Cellular Macromolecules
Plasma membrane proteins
and phosphotipids
Free fatty acids and
carbohydrates
Mechanisms for Antioxidant
Repletion
• Secretion by epithelial ceHs
• Transport from plasma
• Reduction of oxidized ascorbate
V J
Secondary Oxidation Products
Oxidized proteins
Aldehydes
Ozonized cholesterol species
Lipid Peroxides
Eicosanoids and PAF
Hyaluronan Fragments
Cellular injury
Cellular signaling
Mechanisms for Reaction
Product Removal
* Quenching reactions by ELF
ant ioxidant sand proteins
* Non-enzymstic reactions with
cellular aotioxidaots
• Metabolism by cellular GST/WQOl
• Receptor-mediated uptake by
macrophsges J
Contents of this figure not discussed in Section 5.2 will be discussed in Section 5.3. Clara cell secretory protein, CCSP; Surfactant
Protein-A, SP-A; Platelet activating factor, PAF.
Figure 5-8 Details of the Os interaction with the airway ELF to form secondary
oxidation products. Ozone will react with components of the ELF to
produce reaction products that may lead to cellular injury and cell
signaling as discussed in Section 5.3.
1
2
3
4
5
6
1
8
9
10
11
underlying tissues and to prevent penetration of O3 deeper into the lung. The formation of
secondary oxidation products is likely related to the concentration of antioxidants present
and the quenching ability of the lining fluid. Mechanisms are present to replenish the
antioxidant substrate pools as well as to remove secondary reaction products from tissue
interactions. Important differences exist in the reaction rates for O3 and these ELF
biomolecules and the reactivity of the resulting products. Overall, studies suggest that UA
and AH2 are more reactive with O3 than GSH, proteins, or lipids. In addition to contri-
buting to the driving force for O3 uptake, formation of secondary oxidation products may
lead to increased cellular injury and cell signaling (discussed in Section 5.3). Studies
indicate that the antioxidants might be participating in reactions where the resulting
secondary oxidation products might penetrate into the tissue layer and cause injury.
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5.3 Possible Pathways/Modes of Action
5.3.1 Introduction
1 Mode of action refers to a sequence of key events and processes which result in a given
2 toxic effect (U.S. EPA. 2005). Elucidation of mechanisms provides a more detailed
3 understanding of these key events and processes (U.S. EPA. 2005). Moreover, toxicity
4 pathways describe the processes by which perturbation of normal biological processes
5 produce changes sufficient to lead to cell injury and subsequent events such as adverse
6 health effects (U.S. EPA. 2009f). The purpose of this section of Chapter 5 is to describe
7 the key events and toxicity pathways which contribute to health effects resulting from
8 short-term and long-term exposures to O3. The extensive research carried out over
9 several decades in humans and in laboratory animals has yielded numerous studies on
10 mechanisms by which O3 exerts its effects. This section will discuss some of the
11 representative studies with particular emphasis on studies published since the 2006 O3
12 AQCD and on studies in humans which inform biological mechanisms underlying
13 responses to O3.
14 It is well-appreciated that secondary oxidation products, which are formed as a result of
15 O3 exposure, initiate numerous responses at the cellular, tissue and whole organ level of
16 the respiratory system. These responses include the activation of neural reflexes,
17 initiation of inflammation, alteration of epithelial barrier function, sensitization of
18 bronchial smooth muscle, modification of innate/adaptive immunity and airways
19 remodeling, as will be discussed below. Exposure to O3 also may result in effects on
20 other organ systems such as the cardiovascular, central nervous, hepatic and reproductive
21 systems. It is unlikely that lipid ozonides and other secondary oxidation products, which
22 are bioactive and cytotoxic in the respiratory system, gain access to the vascular space
23 (Chuang et al.. 2009). However the inhalation of O3 may result in systemic oxidative
24 stress. The following subsections describe the current understanding of potential
25 pathways and modes of action responsible for the pulmonary and extrapulmonary effects
26 of O3 exposure.
5.3.2 Activation of Neural Reflexes
27 Acute O3 exposure results in reversible effects on lung function parameters through
28 activation of neural reflexes. The involvement of bronchial C-fibers, a type of nociceptive
29 sensory nerve, has been demonstrated in dogs exposed through an endotracheal tube to 2-
30 3 ppm O3 for 20-70 minutes (Coleridge et al.. 1993; Schelegle et al.. 1993). This vagal
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1 afferent pathway was found to be responsible for O3-mediated rapid shallow breathing
2 and other changes in respiratory mechanics in O3-exposed dogs (Schelegle etal.. 1993).
3 Ozone also triggers neural reflexes which stimulate the autonomic nervous system and
4 alter electrophysiologic responses of the heart. For example, bradycardia, altered HRV
5 and arrhythmia have been demonstrated in rodents exposed to 0.1-0.6 ppm O3 (Hamade
6 and Tankersley. 2009; Watkinson et al.. 2001; Aritoetal.. 1990). Another effect is
7 hypothermia, which in rodents occurred subsequent to the activation of neural reflexes
8 involoving the parasympathetic nervous system (Watkinson et al.. 2001). Vagal afferent
9 pathways originating in the respiratory tract may also be responsible for O3-mediated
10 activation of nucleus tractus solitarius neurons which resulted in neuronal activation in
11 stress-responsive regions of the central nervous system (CNS) (rats, 0.5-2.0 ppm O3 for
12 1.5-120hours) (Gackiere etal.. 2011).
13 Recent studies in animals provide new information regarding the effects of O3 on reflex
14 responses mediated by bronchopulmonary C-fibers. In ex vivo mouse lungs, O3 exposure
15 selectively activated a subset of C-fiber receptors which are TRPA1 ion channels (Taylor-
16 Clark and Undem. 2010). TRPA1 ion channels are members of the TRP family of ion
17 channels, which are known to mediate the responses of sensory neurons to inflammatory
18 mediators (Caceres et al.. 2009). In addition to TRPA1 ion channels possibly playing a
19 key role in O3-induced decrements in pulmonary function, they may mediate allergic
20 asthma (Caceres et al.. 2009). Activation of TRPA1 ion channels following O3 exposure
21 is likely initiated by secondary oxidation products such as aldehydes and prostaglandins
22 (Taylor-Clark and Undem. 2010) through covalent modification of cysteine and lysine
23 residues (Trevisani et al.. 2007). Ozonation of unsaturated fatty acids in the ELF was
24 found to result in the generation of aldehydes (Frampton et al.. 1999) such as
25 4-hydroxynonenal and 4-oxononenal (Taylor-Clark et al.. 2008; Trevisani et al.. 2007). 4-
26 oxononenal is a stronger electrophile than 4-hydroxynonenal and exhibits greater potency
27 towards the TRPA1 channels (Taylor-Clark et al.. 2008). (Trevisani etal.. 2007). In
28 addition, PGE2 is known to sensitize TRPA1 channels (Bang et al.. 2007).
29 In exercising humans, the response to O3 (500 ppb for 2 h) was characterized by
30 substernal discomfort, especially on deep inspiration, accompanied by involuntary
31 truncation of inspiration (Hazucha et al.. 1989). This led to decreased inspiratory capacity
32 and to decreased forced vital capacity (FVC) and forced expiratory volume in one second
33 (FEVi), as measured by spirometry. These changes, which occurred during O3 exposure,
34 were accompanied by decreased VT and increased respiratory frequency in human
35 subjects. Spirometric changes in FEVi and FVC were not due to changes in respiratory
36 muscle strength (Hazucha et al.. 1989). In addition, parasympathetic involvement in the
37 O3-mediated decreases in lung volume was minimal (Mudway and Kelly. 2000). since
3 8 changes in FVC or symptoms were not modified by treatment with bronchodilators such
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1 as atropine in exercising human subjects exposed to 400 ppb O3 for 0.5 hour (Beckett et
2 al.. 1985). However, the loss of vital capacity was reversible with intravenous
3 administration of the rapid-acting opioid agonist, sufentanyl, in exercising human
4 subjects exposed to 420 ppb O3 for 2 hours, which indicated the involvement of opioid
5 receptor-containing nerve fibers and/or more central neurons (Passannante et al., 1998).
6 The effects of sufentanyl may be attributed to blocking C-fiber stimulation by O3 since
7 activation of opioid receptors downregulated C-fiber function (Belvisi et al.. 1992). Thus,
8 nociceptive sensory nerves, presumably bronchial C-fibers, are responsible for O3-
9 mediated responses in humans (Passannante et al., 1998). This vagal afferent pathway is
10 responsible for pain-related symptoms and inhibition of maximal inspiration in humans
11 (Hazucha et al.. 1989).
12 There is some evidence that eicosanoids (see Section 5.3.3) play a role in the neural
13 reflex since cyclooxygenase inhibition with indomethacin (Alexis et al.. 2000; Schelegle
14 et al.. 1987) or ibuprofen, which also blocks some lipoxygenase activity (Hazucha et al.,
15 1996). before exposure to O3 significantly blunted the spirometric responses. These
16 studies involved exposures of 1-2 hours to 350-400 ppb O3 in exercising human subjects.
17 In the latter study, ibuprofen treatment resulted in measurable decreases in BALF levels
18 of PGE2 and TXB2 at 1-hour postexposure (Hazucha et al.. 1996). Although an earlier
19 study demonstrated that PGE2 stimulated bronchial C-fibers (Coleridge et al.. 1993;
20 Coleridge etal.. 1976) and suggested that PGE2 mediated O3-induced decreases in
21 pulmonary function, no correlation was observed between the degree of ibuprofen-
22 induced inhibition of BALF PGE2 levels and blunting of the spirometric response to O3
23 (Hazucha et al.. 1996). These results point to the involvement of a lipoxygenase product.
24 Further, as noted above, PGE2 may play a role in the neural reflex by sensitizing TRPA1
25 channels. A recent study in exercising human subjects exposed for 1 hour to 350 ppb O3
26 also provided evidence that arachidonic acid metabolites, as well as oxidative stress,
27 contribute to human responsiveness to O3 (Alfaro et al.. 2007).
28 In addition to the spirometric changes, mild airways obstruction occurred in exercising
29 humans during O3 exposure (500 ppb for 2 hours) (Hazucha et al.. 1989). This pulmonary
30 function decrement is generally measured as specific airway resistance (sRaw) which is
31 the product of airway resistance and thoracic gas volume. In several studies involving
32 exercising human subjects exposed for 1-4 hours to 200-300 ppb O3, changes in sRaw
33 correlated with changes in inflammatory and injury endpoints measured 18-hours
34 postexposure, but did not follow the same time course or change to the same degree as
35 spirometric changes (i.e. FEVi, FVC) measured during exposure (Balmes etal.. 1996;
36 Ariset al.. 1993; Schelegle et al.. 1991). In addition, a small but persistent increase in
37 airway resistance associated with narrowing of small peripheral airways (measured as
38 changes in isovolumetric FEF25_75) was demonstrated in O3-exposed human subjects (350
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1 ppb for 130 minutes with exercise) (Weinmann et al.. 1995a; Weinmann et al.. 1995b). A
2 similar study (400 ppb O3 for 2 hours in exercising human subjects) found decreases in
3 FEF 25.75 concomitant with increases in residual volume, which is suggestive of small
4 airways dysfunction (Kreit et al.. 1989). In separate studies, a statistically significant
5 increase in residual volume (500 ppb for 2 hours) (Hazucha et al.. 1989) and a
6 statistically significant decrease in FEF25-75 (160 ppb for 7.6 hours) (Horstman et al..
7 1995) were observed following O3 exposure in exercising human subjects, providing
8 further support for an O3-induced effect on small airways.
9 Mechanisms underlying this rapid increase in airway resistance following O3 exposure
10 are incompletely understood. Pretreatment with atropine decreased baseline sRaw and
11 prevented O3-induced increases in sRaw in exercising human subjects (400 ppb for 0.5
12 hours) (Beckett et al.. 1985). indicating the involvement of muscarinic cholinergic
13 receptors of the parasympathetic nervous system. Interestingly, atropine pretreatment
14 partially blocked the decrease in FEVi, but had no effect on the decrease in FVC,
15 breathing rate, tidal volume or respiratory symptoms (Beckett et al.. 1985). Using a (3-
16 adrenergic agonist, it was shown that smooth muscle contraction, not increased airway
17 mucus secretion, was responsible for O3-induced increases in airway resistance (Beckett
18 etal.. 1985). Thus, pulmonary function decrements measured as FEVi may reflect both
19 restrictive (such as decreased inspiratory capacity) and obstructive (such as
20 bronchoconstriction) type changes in airway responses. This is consistent with
21 McDonnell et al. (1983) who observed a relatively strong correlation between sRaw and
22 FEVi (r=-0.31, p=0.001) and a far weaker correlation between sRaw and FVC (r=-0.16,
23 p=0.10) in exercising human subjects exposed for 2.5 hours to 120-400 ppb O3.
24 Furthermore, tachykinins may contribute to O3-mediated increases in airway resistance.
25 In addition to stimulating CNS reflexes, bronchopulmonary C-fibers mediate local axon
26 responses by releasing neuropeptides such as substance P (SP), neurokinin (NK) A and
27 calcitonin gene-related peptide (CGRP). Tachykinins bind to NK receptors resulting in
28 responses such as bronchoconstriction. Recent studies in animals demonstrated that NK-1
29 receptor blockade had no effect on O3-stimulated physiologic responses such as VT and
30 fB in rats over the 8 hour exposure to 1 ppm O3 (Oslund et al., 2008). However, SP and
31 NK receptors contributed to vagally-mediated bronchoconstriction in guinea pigs 3 days
32 after a single 4-hour exposure to 2 ppm O3 (Verhein et al., 2011). In one human study in
33 which bronchial biopsies were performed and studied by immunohistochemistry, SP was
34 substantially diminished in submucosal sensory nerves 6 hours following O3 exposure
35 (200 ppb for 2 hours with exercise) (Krishna et al.. 1997). A statistically significant
36 correlation was observed between loss of SP immunoreactivity from neurons in the
37 bronchial mucosa and changes in FEVi measured 1-hour postexposure (Krishna et al..
38 1997). Another study found that SP was increased in lavage fluid of human subjects
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1 immediately after O3 challenge (250 ppb for 1 hour with exercise) (Hazbun et al.. 1993).
2 These results provide evidence that the increased airway resistance observed following
3 O3 exposure is due to vagally-mediated responses and possibly by local axon reflex
4 responses through bronchopulmonary C-fiber-mediated release of SP.
5.3.3 Initiation of inflammation
5 As described previously (5.2.3), O3 reacts with components of the ELF and cellular
6 membranes resulting in the generation of secondary oxidation products. Higher
7 concentrations of these products may directly injure respiratory tract epithelium. Lower
8 concentrations may initiate cellular responses including cytokine generation, adhesion
9 molecule expression and modification of tight junctions leading to inflammation and
10 increased permeability across airway epithelium (Section 5.3.4) (Dahl et al.. 2007;
11 Mudway and Kelly. 2000). Subsequent airways remodeling may also occur (Section
12 5.3.7) (Mudwav and Kelly. 2000).
13 An important hallmark of acute O3 exposure in humans and animals is neutrophilic
14 airways inflammation. Although neutrophil influx into nasal airways has been
15 demonstrated in exercising human subjects (400 ppb O3, 2 hours) (Graham and Koren.
16 1990). most studies of neutrophil influx have focused on the lower airways (Hazucha et
17 al.. 1996; Aris et al.. 1993). The time course of this response in the lower airways and its
18 resolution was slower than that of the decrements in pulmonary function in exercising
19 human subjects exposed for 2 hours to 500 ppb O3 (Hazucha et al., 1996). In several
20 studies, airways neutrophilia was observable within 1-2 hours, peaked at 4-6 hours and
21 was returning to baseline levels at 24 hours following exposure of 1-2 hours to 300-400
22 ppb O3 in exercising humans (Devlin et al.. 1991; Schelegle et al.. 1991). Since the influx
23 and persistence of neutrophils in airways following O3 exposure correlated with the
24 temporal profile of epithelial injury (guinea pigs, 0.26-1 ppm O3 72 hours) (Hu et al..
25 1982). neutrophils were probably injurious. However, neutrophils have also been shown
26 to contribute to repair of O3-injured epithelium in rats exposed for 8 hours to 1 ppm O3
27 possibly by removing necrotic epithelial cells (Mudway and Kelly. 2000; Vesely et al..
28 1999). Nonetheless, the degree of airways inflammation due to O3 is thought to have
29 more important long-term consequences than the more quickly resolving changes in
30 pulmonary function since airways inflammation is often accompanied by tissue injury
31 (Balmes et al.. 1996).
32 Ozone exposure results in alterations in other airways inflammatory cells besides
33 neutrophils, including lymphocytes, macrophages, monocytes and mast cells. Influx of
34 some of these cells accounts for the later (i.e. 18-20 hours) phase of inflammation
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1 following O3 exposure. Numbers of lymphocytes and total cells in BALF were decreased
2 early after O3 exposure in exercising humans exposed for 2 hours to 200 ppb O3, which
3 preceded the neutrophil influx (Mudway and Kelly. 2000; Blomberg et al.. 1999; Krishna
4 et al.. 1997). The decrease in total cells was thought to reflect decreases in macrophages,
5 although it was not clear whether the cells were necrotic or whether membrane adhesive
6 properties were altered making them more difficult to obtain by lavage (Mudway and
7 Kelly. 2000; Blomberg et al.. 1999; Mudwav et al.. 1999b; Frampton et al.. 1997a;
8 Pearson and Bhalla. 1997). A recent study in exercising human subjects exposed for 6.6
9 hours to 80 ppb O3 demonstrated an increase in numbers of sputum monocytes and
10 dendritic-like cells with increased expression of innate immune surface proteins and
11 antigen presentation markers (Peden. 2011; Alexis et al.. 2010) (see Section 6.2.3.1). An
12 increase in submucosal mast cells was observed 1.5 hours after a 2 hour-exposure to 200
13 ppb O3 (Blomberg et al., 1999) and an increase in BAL mast cell number was observed
14 18 hours after a 4-hour exposure to 220 ppb O3 exposure in exercising human subjects
15 (Frampton et al.. 1997a). Mast cells may play an important role in mediating neutrophil
16 influx since they are an important source of several pro-inflammatory cytokines and since
17 their influx preceded that of neutrophils in exercising human subjects exposed for 2 hours
18 to 200 ppb O3 (Stenfors et al.. 2002; Blomberg et al.. 1999). Further, a study using mast
19 cell-deficient mice demonstrated decreased neutrophilic inflammation in response to O3
20 (1.75 ppm, 3 hours) compared with wild type mice (Kleeberger et al., 1993). Influx of
21 these inflammatory cell types in the lung is indicative of O3-mediated activation of innate
22 immunity as will be discussed in Section 5.3.6.
23 Much is known about the cellular and molecular signals involved in inflammatory
24 responses to O3 exposure (U.S. EPA. 2006b). Eicosanoids are one class of secondary
25 oxidation products which may be formed rapidly following O3 exposure and which may
26 mediate inflammation. In addition, secondary reaction products may stimulate
27 macrophages to produce cytokines such as IL-1, IL-6 and TNF-a which in turn activate
28 IL-8 production by epithelial cells. Although IL-8 has been proposed to play a role in
29 neutrophil chemotaxis, measurements of IL-8 in BALF from humans exposed to O3
30 found increases that were too late to account for this effect (Mudway and Kelly. 2000).
31 The time-course profiles of PGE2 and IL-6 responses suggest that they may play a role in
32 neutrophil chemotaxis in humans (Mudway and Kelly. 2000). However, pretreatment
33 with ibuprofen attenuated O3-induced increases in BALF PGE2 levels, but had no effect
34 on neutrophilia in exercising human subjects exposed for 2 hour to 400 ppb O3 (Hazucha
35 etal.. 1996).
36 One set of studies in humans focused on the earliest phase of airways inflammation (1-2
37 hours following exposure). Exercising subjects were exposed to 200 ppb O3 for 2 hours
38 and bronchial biopsy tissues were obtained 1.5 and 6 hours after exposure (Bosson et al..
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1 2009: Bosson et al.. 2003; Stenfors et al.. 2002; Blomberg et al.. 1999). Results
2 demonstrated upregulation of vascular endothelial adhesion molecules P-selectin and
3 ICAM-1 at both 1.5 and 6 hours (Stenfors et al.. 2002; Blomberg et al.. 1999).
4 Submucosal mast cell numbers were increased at 1.5 hours in the biopsy samples without
5 an accompanying increase in neutrophil number (Blomberg et al.. 1999). Pronounced
6 neutrophil infiltration was observed at 6 hours in the bronchial mucosa (Stenfors et al..
7 2002). Surprisingly, suppression of the NF-KB and AP-1 pathways at 1.5 hours and a lack
8 of increased IL-8 at 1.5 or 6 hours in bronchial epithelium was observed (Bosson et al..
9 2009). The authors suggested that vascular endothelial adhesion molecules, rather than
10 redox sensitive transcription factors, play key roles in early neutrophil recruitment in
11 response to O3.
12 Increases in markers of inflammation occurred to a comparable degree in exercising
13 human subjects with mild (least sensitive) and more remarkable (more sensitive)
14 spirometric responses to O3 (200 ppb, 4 hours) (Balmes et al.. 1996). Two other studies
15 using similar protocols (200 ppb for 4 hours and 300 ppb for 1 hour) found that acute
16 spirometric changes were not positively correlated with cellular and biochemical
17 indicators of inflammation (Aris et al.. 1993; Schelegle et al.. 1991). However
18 inflammation was correlated with changes in sRaw (Balmes et al.. 1996). In another
19 study, pretreatment with ibuprofen had no effect on neutrophilia although it blunted the
20 spirometric response in exercising human subjects exposed for 2 hours to 400 ppb O3
21 (Hazucha et al.. 1996). Taken together, results from these studies indicate different
22 mechanisms underlying the spirometric and inflammatory responses to O3.
23 A common mechanism underlying both inflammation and impaired pulmonary function
24 was suggested by Krishna et al. (1997). This study, conducted in exercising humans
25 exposed to 200 ppb O3 for 2 hours, demonstrated a correlation between loss of SP
26 immunoreactivity from neurons in the bronchial mucosa and numbers of neutrophils and
27 epithelial cells (shed epithelial cells are an index of injury) in the BALF 64iours
28 postexposure. Furthermore, the loss of SP immunoreactivity was correlated with the
29 observed changes in FEVi. Another study found that SP was increased in lavage fluid of
30 exercising human subjects immediately after O3 challenge (250 ppb, 1 hour) (Hazbun et
31 al.. 1993). SP is a neuropeptide released by sensory nerves which mediates neurogenic
32 edema and bronchoconstriction (Krishna et al.. 1997). Taken together, these findings
33 suggest that O3-mediated stimulation of sensory nerves which leads to activation of
34 central and local axon reflexes is s a common effector pathway leading to impaired
35 pulmonary function and inflammation.
36 Studies in animal models have confirmed many of these findings and provided evidence
37 for additional mechanisms involved in O3-induced inflammation. A study in mice (2 ppm
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1 O3, 3 hours) demonstrated that PAF may be important in neutrophil chemotaxis
2 (Longphre etal.. 1999). while ICAM-1 and macrophage inflammatory protein-2 (MIP-2),
3 the rodent IL-8 homologue, have been implicated in a rat model (1 ppm O3, 3 hours)
4 (Bhalla and Gupta. 2000). Key roles for CXCR2, a receptor for keratinocyte-derived
5 chemokine (KC) and MIP-2, and for IL-6 in O3-mediated neutrophil influx were
6 demonstrated in mice (1 ppm O3, 3 hours) (Johnston et al.. 2005a: Johnston et al..
7 2005b). Activation of JNK and p38 pathways and cathepsin-S were also found to be
8 important in this response (3 ppm O3, 3 hours) (Williams et al.. 2009a: Williams et al..
9 2008b; Williams et al.. 2007a). Matrix metalloproteinase-9 (MMP-9) protected against
10 O3-induced airways inflammation and injury in mice (0.3 ppm O3, 6-72 hours) (Yoon et
11 al.. 2007). Interleukin-10 (IL-10) was also found to be protective since IL-10 deficient
12 mice responded to O3 exposure (0.3 ppm, 24-72 hours) with enhanced numbers of BAL
13 neutrophils, enhanced NF-KB activation and MIP-2 levels compared with IL-10 sufficient
14 mice (Backus etal.. 2010).
15 In addition, lung epithelial cells may release ATP in response to O3 exposure (Ahmad et
16 al.. 2005). ATP and its metabolites (catalyzed by ecto-enzymes) can bind to cellular
17 purinergic receptors resulting in activation of cell signaling pathways (Picher et al..
18 2004). One such metabolite, adenine, is capable of undergoing oxidation leading to the
19 formation of UA which, if present in high concentrations, could activate inflammasomes
20 and result in caspase 1 activation and the maturation and secretion of IL-1(3 and IL-18
21 (Dostert et al.. 2008). A recent study in exercising human subjects exposed for 2 hours to
22 400 ppb O3 demonstrated a correlation between ATP metabolites and inflammatory
23 markers (Esther et al.. 2011). which provides some support for this mechanism.
24 Several recent studies have focused on the role of toll-like receptor (TLR) and its related
25 adaptor protein MyD88 in mediating O3-induced neutrophilia. While Hollingsworth et al.
26 (2004) demonstrated airways neutrophilia which was TLR4-independent following acute
27 (2 ppm, 3 hours) and subchronic (0.3 ppm, 72 hours) O3 exposure in a mouse model,
28 Williams et al. (2007b) found that MyD88 was important in mediating O3-induced
29 neutrophilia in mice (3 ppm, 3 hours), with TLR4 and TLR2 contributing to the speed of
30 the response. Moreover, MyD88, TLR2 and TLR4 contributed to inflammatory gene
31 expression in this model and O3 upregulated MyD88, TLR4 and TLR4 gene expression
32 (Williams et al.. 2007a)
33 Hyaluronan was found to mediate a later phase (24 hours) of O3-induced inflammation in
34 mice (Garantziotis et al.. 2010; Garantziotis et al.. 2009). Hyaluronan is an extracellular
3 5 matrix component which is normally found in the ELF as a large polymer. Exposure to
36 2 ppm O3 for 3 hours resulted in elevated levels of soluble low molecular weight
37 hyaluronan in the BALF 24-hours postexposure (Garantziotis et al.. 2010; Garantziotis et
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1 al.. 2009). Ozone may have caused the depolymerization of hyaluronan to soluble
2 fragments which are known to be endogenous ligands of the CD44 receptor and TLR4 in
3 the macrophage (Jiang etal. 2005). Binding of hyaluronan fragments to the CD44
4 receptor activates hyaluronan clearance, while binding to TLR4 results in signaling
5 through MyD88 to produce chemokines that stimulate the influx of inflammatory cells
6 (Jiang et al.. 2005). Activation of NF-KB occurred in both airway epithelia and alveolar
7 macrophages 24-hours postexposure to O3. Increases in BALF pro-inflammatory factors
8 KC, IL-1|3, MCP-1, TNF-a and IL-6 observed 24 hours following O3 exposure were
9 found to be partially dependent on TLR4 (Garantziotis et al., 2010) while increases in
10 BAL inflammatory cells, which consisted mainly of macrophages, were dependent on
11 CD44 (Garantziotis et al.. 2009). BAL inflammatory cells number and injury markers
12 following O3 exposure were similar in wild-type and TLR4-deficient animals
13 (Garantziotis et al.. 2010).
14 Since exposure to O3 leads to airways inflammation characterized by neutrophilia, and
15 since neutrophil-derived oxidants often scavenge ELF antioxidants, concentrations of
16 ELF antioxidants have been examined during airways neutrophilia (Long etal.. 2001;
17 Gunnison and Hatch. 1999; Mudway et al.. 1999b). In exercising humans exposed to 200
18 ppb O3 for 2 hours, UA, GSH and a-TOH levels remained unchanged in BALF 6-hours
19 postexposure while AH2 was decreased significantly in both BALF and plasma (Mudway
20 et al.. 1999b). A second study involving the same protocol reported a loss of AH2 from
21 bronchial wash fluid and BALF, representing proximal and distal airway ELF
22 respectively, as well as an increase in oxidized GSH in both compartments (Mudway et
23 al.. 2001). No change was observed in ELF UA levels in response to O3 (Mudway et al..
24 2001). Further, O3 exposure (0.8 ppm, 4 hours) in female rats resulted in a 50% decrease
25 in BALF AH2 immediately postexposure (Gunnison and Hatch. 1999). These studies
26 suggested a role for AH2 and GSH in protecting against the oxidative stress associated
27 with inflammation.
5.3.4 Alteration of epithelial barrier function
28 Following O3 exposure, injury and inflammation can lead to altered airway barrier
29 function. Histologic analysis has demonstrated damage to tight junctions between
30 epithelial cells, suggesting an increase in epithelial permeability. In addition, the presence
31 of shed epithelial cells in the BALF and increased epithelial permeability, which is
32 measured as the flux of small solutes, have been observed and are indicative of epithelial
33 injury. Increases in vascular permeability, as measured by BALF protein and albumin,
34 have also been demonstrated (Costa et al.. 1985; Hu et al.. 1982).
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1 An early study in sheep measured changes in airway permeability as the flux of inhaled
2 radiolabeled histamine into the plasma (Abraham et al.. 1984). Exposure of sheep to 0.5
3 ppm O3 for 2 hours via an endotracheal tube resulted in an increased rate of histamine
4 appearance in the plasma at 1 day postexposure. Subsequently, numerous studies have
5 measured epithelial permeability as the flux of the small solute 99mTcDTPA which was
6 introduced into the air spaces in different regions of the respiratory tract. Increased
7 pulmonary epithelial permeability, measured as the clearance of 99mTc-DTPA, was
8 demonstrated in humans 1-2 hours following a 2-hour exposure to 400 ppb O3 while
9 exercising moderately (Kehrl et al., 1987). Another study in human subjects found
10 increased epithelial permeability 19-hours postexposure to 240 ppb O3 for 130 minutes
11 while exercising (Foster and Stetkiewicz. 1996). Increased bronchial permeability was
12 also observed in dogs 1-day postexposure (0.4 ppm O3 by endotracheal tube for 6 hours)
13 and did not resolve for several days (Foster and Freed. 1999).
14 A role for tachykinins in mediating airway epithelial injury and decreased barrier function
15 has been suggested. Nishiyama et al. (1998) demonstrated that capsaicin, which depletes
16 nerve fibers of substance P, blocked the O3-induced increase in permeability of guinea
17 pig tracheal mucosa (0.5-3 ppm O3, 0.5 hours). Pretreatment with propranolol or atropine
18 failed to inhibit this response, suggesting that adrenergic and cholinergic pathways were
19 not involved. In another study, tachykinins working through NK-1 and CGRP receptors
20 were found to contribute to airway epithelial injury in O3-exposed rats (1 ppm, 8 hours)
21 (Oslund et al.. 2009. 2008).
22 Kleeberger et al. (2000) evaluated genetic susceptibility to O3-induced altered barrier
23 function in recombinant inbred strains of mice. Lung hyperpermeability, measured as
24 BALF protein, was evaluated 72 hours after exposure to 0.3 ppm O3 and found to be
25 associated with a functioning TLR4 gene. This study concluded that Tlr4 was a strong
26 candidate gene for susceptibility to hyperpermeability in response to O3 (Kleeberger et
27 al.. 2000). A subsequent study by these same investigators found that Tlr4 modulated
28 Nos2 mRNA levels and suggested that the gene product of Nos2, iNOS, plays an
29 important role in O2-induced lung hyperpermeability (0.3 ppm, 72 hours) (Kleeberger et
30 al., 2001). More recently, HSP70 was identified as part of the TLR4 signaling pathway
31 (0.3 ppm, 6-72 hours) (Bauer et al.. 2011).
32 Antioxidants have been shown to confer resistance to O3-induced injury. In a recent
33 study, lung hyperpermeability in response to O3 (0.3 ppm, 48 hours) was unexpectedly
34 reduced in mice deficient in the glutamate-cysteine ligase modifier subunit gene
35 compared with sufficient mice (Johansson et al., 2010). Since the lungs of these mice
36 exhibited 70% glutathione depletion, protection against O3-induced injury was
37 unexpected (Johansson et al.. 2010). However it was found that several other antioxidant
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1 defenses, including metallothionein, were upregulated in response to O3 to a greater
2 degree in the glutathione-deficient mice compared with sufficient mice (Johansson et al..
3 2010). The authors suggested that resistance to O3-induced lung injury was due to
4 compensatory augmentation of antioxidant defenses (Johansson et al.. 2010). Antioxidant
5 effects have also been attributed to Clara cell secretory protein (CCSP) and surfactant
6 protein A (SP-A). CCSP was found to modulate the susceptibility of airway epithelium to
7 injury in mice exposed to O3 (0.2 or 1 ppm for 8 hours) by an unknown mechanism
8 (Plopper et al.. 2006). SP-A protected against O3-induced airways inflammation and
9 injury in mice (2 ppm, 3 hours), possibly by acting as a sacrificial substrate (Hague et al..
10 2007).
11 Increased epithelial permeability has been proposed to play a role in allergic sensitization
12 (Matsumura. 1970). in activation of neural reflexes and in stimulation of smooth muscle
13 receptors (Dimeo et al.. 1981). Abraham et al. (1984) reported a correlation between
14 airway permeability and airways hyperresponsiveness (AHR) in O3-exposed sheep.
15 However a recent study in human subjects exposed to 220 ppb O3 for 135 minutes while
16 exercising did not find a relationship between O3-induced changes in airway permeability
17 and AHR (Oue et al.).
5.3.5 Sensitization of bronchial smooth muscle
18 Bronchial reactivity is generally determined in terms of a response to a challenge agent.
19 Non-specific bronchial reactivity in humans is assessed by measuring the effect of
20 inhaling increasing concentrations of a bronchoconstrictive drug on lung mechanics
21 (sRaw or FEVi). Methacholine is most commonly employed but histamine and other
22 agents are also used. Specific bronchial reactivity is assessed by measuring effects in
23 response to an inhaled allergen in individuals (or animals) already sensitized to that
24 allergen. An increase in sRaw in response to non-specific or specific challenge agents
25 indicates AHR.
26 In addition to causing mild airway obstruction as discussed above, acute O3 exposure
27 results in reversible increases in bronchial reactivity by mechanisms which are not well
28 understood. In one study, bronchial reactivity of healthy subjects was significantly
29 increased 19-hours postexposure to O3 (120-240 ppb O3 for 2 hours with intermittent
30 exercise) (Foster et al.. 2000). These effects may be more significant in human subjects
31 with already compromised airways (Section 5.4.2.2).
32 Ozone may sensitize bronchial smooth muscle to stimulation through a direct effect on
33 smooth muscle or through effects on the sensory nerves in the epithelium or on the motor
34 nerves innervating the smooth muscle (O'Bvrne et al.. 1984; O'Byrne etal.. 1983;
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1 Holtzman et al., 1979). It is also recognized that increased bronchial reactivity can be
2 both a rapidly occurring and a persistent response to O3 (Foster and Freed. 1999).
3 Tachykinins and secondary oxidation products of O3 have been proposed as mediators of
4 the early response and inflammation-derived products have been proposed as mediators
5 of the later response (Foster and Freed, 1999).
6 Ozone-induced increases in epithelial permeability, which could improve access of
7 agonist to smooth muscle receptors, may be one mechanism of sensitization through a
8 direct effect on bronchial smooth muscle (Holtzman et al., 1979). As noted above, a
9 correlation between airway permeability and AHR has been reported in O3-exposed
10 sheep (Abraham et al., 1984) but not in O3-exposed human subjects (Oue et al.).
11 Neurally-mediated sensitization has been demonstrated. In human subjects exposed for 2
12 hours to 600 ppb O3 while exercising, pretreatment with atropine inhibited O3-induced
13 AHR, suggesting the involvement of cholinergic postganglionic pathways (Holtzman et
14 al.. 1979). Animal studies have demonstrated that O3-induced AHR involved vagally-
15 mediated responses (rabbits, 0.2 ppm O3, 72 hours) (Freed et al.. 1996) and local axon
16 reflex responses through bronchopulmonary C-fiber-mediated release of SP (guinea pigs,
17 0.8 ppm O3, 2 hours) (Joadetal.. 1996). Further, pretreatment with capsaicin to deplete
18 nerve fibers of SP blocked O3-mediated AHR (guinea pigs, 1-2 ppm O3, 2-2.25 hours)
19 (Tepper et al.. 1993). Other investigators demonstrated that SP released from airway
20 nociceptive neurons in ferrets contributed to O3-induced AHR (2 ppm O3, 3 hours) (Wu
21 et al.. 2008b: Wu et al.. 2003).
22 Some evidence suggests the involvement of arachidonic acid metabolites and neutrophils
23 in mediating O3-induced AHR (Seltzer et al., 1986; Fabbrietal., 1985). Increased BAL
24 neutrophils and cyclooxygenase products were found in one study demonstrating AHR in
25 exercising humans (600 ppb for 2 hours) immediately postexposure to (Seltzer et al.,
26 1986). Another study found that ibuprofen pretreatment had no effect on AHR in
27 exercising humans following exposure to 400 ppb O3 for 2 hours, although spirometric
28 responses were blunted (Hazucha et al.. 1996). This study indicated that the arachidonic
29 acid metabolites whose generation was blocked by ibuprofen, (i.e. prostaglandins,
30 thromboxanes and some leukotrienes) did not play a role in AHR. Experiments in dogs
31 exposed for 2 hours to 2.1 ppm O3 demonstrated a close correlation between O3-induced
32 AHR and airways neutrophilic inflammation measured in tissue biopsies (Holtzman et
33 al., 1983). Furthermore, the increased AHR observed in dogs following O3 exposure (3
34 ppm, 2 hours) was inhibited by neutrophil depletion (O'Byrne et al.. 1983) and by pre-
35 treatment with inhibitors of arachidonic acid metabolism. In one of these studies,
36 indomethacin pre-treatment did not prevent airways neutrophilia in response to O3 (3
37 ppm, 2 hours) providing evidence that the subset of arachidonic acid metabolites whose
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1 generation was inhibitable by the cyclooxygenase inhibitor indomethacin (i.e.,
2 prostaglandins and thromboxanes) was not responsible for neutrophil influx (O'Byrne et
3 al.. 1984). Taken together, these findings suggest that arachidonic acid metabolites, but
4 probably not prostaglandins or thromboxanes, may be involved in the AHR response
5 following O3 exposure in dogs. Studies probing the role of neutrophils in mediating the
6 AHR response have provided inconsistent results (Al-Hegelan et al. 2011).
7 Evidence for cytokine and chemokine involvement in the AHR response to O3 has been
8 described. Some studies have suggested a role for TNF-a (mice, 0.5 and 2 ppm O3, 3
9 hours) (Cho et al.. 2001; Shore et al.. 2001) and IL-1 (mice and ferrets, 2 ppm O3, 3
10 hours) (Wu et al.. 2008b; Park et al.. 2004). The latter study found that SP expression in
11 airway neurons was upregulated by IL-1 which was released in response to O3. Other
12 studies in mice have demonstrated a key role for CXCR2, the chemokine receptor for the
13 neutrophil chemokines KC and MIP-2, but not for IL-6 in O3-mediated AHR ( 1 ppm O3,
14 3 hours) (Johnston et al.. 2005a; Johnston et al.. 2005b). In contrast, CXCR2 and IL-6
15 were both required for neutrophil influx in this model (Johnston et al.. 2005a: Johnston et
16 al.. 2005b). as discussed above. Williams et al. (2008a) demonstrated that the Th2
17 cytokine IL-13 contributed to AHR, as well as to airways neutrophilia, in mice (3 ppm
18 O3, 3 hours).
19 Other studies have focused on the role of TLR4. Hollingsworth et al. (2004) measured
20 AHR, as well as airways neutrophilia, in mice 6 and 24 hours following acute (2 ppm O3
21 for 3 hours) and subchronic (0.3 ppm for 3 days) exposure to O3. TLR4 is a key
22 component of the innate immune system and is responsible for the immediate
23 inflammatory response seen following challenge with endotoxin and other pathogen-
24 associated substances. In this study, a functioning TLR4 was required for the full AHR
25 response following O3 exposure but not for airways neutrophilia (Hollingsworth et al..
26 2004). These findings are complemented by an older study demonstrating that O3 effects
27 on lung hyperpermeability required a functioning TLR4 (mice, 0.3 ppm O3, 72 hours)
28 (Kleeberger et al.. 2000). Williams et al. (2007b) found that TLR2, TLR4 and the TLR
29 adaptor protein MyD88 contributed to AHR in mice (3 ppm O3, 3 hours). Ozone was also
30 found to upregulate MyD88, TLR4 and TLR4 gene expression in this model (Williams et
31 al.. 2007b).
32 A newly recognized mechanistic basis for O3-induced AHR is provided by studies
33 focusing on the role of hyaluronan following O3 exposure in mice (Garantziotis et al..
34 2010; Garantziotis et al.. 2009). Hyaluronan is an extracellular matrix component which
35 is normally found in the ELF as a large polymer. Briefly, TLR4 and CD44 were found to
36 mediate AHR in response to O3 and hyaluronan. Exposure to 2 ppm O3 for 3 hours
3 7 resulted in enhanced AHR and elevated levels of soluble low molecular weight
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1 hyaluronan in the BALF 24-hours postexposure (Garantziotis et al., 2010; Garantziotis et
2 al.. 2009). Ozone may have caused the depolymerization of hyaluronan to soluble
3 fragments which are known to be endogenous ligands of the CD44 receptor and TLR4 in
4 the macrophage (Jiang et al.. 2005). In the two recent studies, O3-induced AHR was
5 attenuated in CD44 and TLR4-deficient mice (Garantziotis et al., 2010; Garantziotis et
6 al.. 2009). Hyaluronan fragment-mediated stimulation of AHR was found to require
7 functioning CD44 receptor and TLR4 (Garantziotis et al., 2010; Garantziotis et al.. 2009).
8 In contrast, high-molecular-weight hyaluronan blocked AHR in response to O3
9 (Garantziotis et al.. 2009). In another study high-molecular-weight hyaluronan enhanced
10 repair of epithelial injury (Jiang et al.. 2005). These studies provide a link between innate
11 immunity and the development of AHR following O3 exposure, and indicate a role for
12 TLR4 in increasing airways responsiveness. While TLR4-dependent responses usually
13 involve activation of NF-KB and the upregulation of proinflammatory factors, the precise
14 mechanisms leading to AHR are unknown (Al-Hegelan et al.. 2011).
15 In guinea pigs, AHR was found to be mediated by different pathways at 1- and 3-days
16 postexposure to a single dose of O3 (2 ppm for 4 hours) (Verhein etal., 2011; Yost et al.,
17 2005). At 1 day, AHR was due to activation of airway parasympathetic nerves rather than
18 to a direct effect on smooth muscle (Yostet al.. 2005). This effect occurred as a result of
19 O3-stimulated release of major basic protein from eosinophils (Yost et al., 2005). Major
20 basic protein is known to block inhibitory M2 muscarinic receptors which normally
21 dampen acetylcholine release from parasympathetic nerves (Yostet al., 2005). The
22 resulting increase in acetylcholine release caused an increase in smooth muscle
23 contraction following O3 exposure (Yostetal., 2005). Eosinophils played a different role
24 3-days postexposure to O3 in guinea pigs (Yost et al.. 2005). Ozone-mediated influx of
25 eosinophils into lung airways resulted in a different population of cells present 3-days
26 postexposure compared to those present at 1 day (Yost et al.. 2005). At this time point,
27 eosinophil-derived major basic protein increased smooth muscle responsiveness to
28 acetylcholine which also contributed to AHR (Yost et al.. 2005). However, the major
29 effect of eosinophils was to protect against vagal hyperreactivity (Yost et al., 2005). The
30 authors suggested that these beneficial effects were due to the production of nerve growth
31 factor (Yost et al., 2005). Further work by these investigators demonstrated a key role for
32 IL-1|3 in mediating AHR 3-days postexposure to O3 (Verhein et al.. 2011). In this study,
33 IL-1|3 increased nerve growth factor and SP which acted through the NK1 receptor to
34 cause vagally-mediated bronchoconstriction (Verhein et al.. 2011). The mechanism by
35 which SP caused acetylcholine release from parasympathetic nerves following O3
36 exposure was not determined (Verhein et al.. 2011). Taken together, the above study
37 results indicate that mechanisms involved in O3-mediated AHR can vary over time
38 postexposure and that eosinophils and SP can play a role. Results of this animal model
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1 may provide some insight into allergic airways disease in humans which is characterized
2 by eosinophilia (Section 5.4.2.2).
5.3.6 Modification of innate/adaptive immune system responses
3 Host defense depends on effective barrier function and on innate immunity and adaptive
4 immunity (Al-Hegelan et al.. 2011). Ozone's effect on barrier function in the airways was
5 discussed above (Section 5.3.4). This section focuses on the mechanisms by which O3
6 impacts innate and adaptive immunity. Both tissue damage and foreign pathogens are
7 triggers for the activation of the innate immune system. This results in the influx of
8 inflammatory cells such as neutrophils, mast cells, basophils, eosinophils, monocytes and
9 dendritic cells and the generation of cytokines such as TNF-a, IL-1, IL-6, KC and IL-17 .
10 Further, innate immunity encompasses the actions of complement and collectins and the
11 phagocytic functions of macrophages, neutrophils and dendritic cells. Airway epithelium
12 also contributes to innate immune responses. Innate immunity is highly dependent on cell
13 signaling networks involving TLR4. Adaptive immunity provides immunologic memory
14 through the actions of B and T cells. Important links between the two systems are
15 provided by dendritic cells and antigen presentation. Recent studies demonstrate that
16 exposure to O3 modifies cells and processes which are required for innate immunity,
17 contributes to innate-adaptive immune system interaction and primes pulmonary immune
18 responses to endotoxin.
19 Ozone exposure of human subjects resulted in recruitment of activated innate immune
20 cells to the airways. Healthy individuals were exposed to 80 ppb O3 for 6.6 hours with
21 intermittent exercise and airways inflammation was characterized in induced sputum 18-
22 hours postexposure (Alexis et al.. 2010). Previous studies demonstrated that induced
23 sputum contains liquid and cellular constituents of the ELF from central conducting
24 airways (Alexis et al.. 200 Ib) and also identified these airways as a site of preferential O3
25 absorption during exercise (Hu et al.. 1994). Ozone exposure resulted in increased
26 numbers of neutrophils, airway monocytes and dendritic-like cells in sputum (Alexis et
27 al.. 2010). In addition, increased expression of cell surface markers characteristic of
28 innate immunity and antigen presentation (i.e. CD-14 and HLA-DR) was demonstrated
29 on airway monocytes (Alexis et al.. 2010). Enhanced antigen presentation contributes to
30 exaggerated T cell responses and promotes Th2 inflammation and an allergic phenotype
31 (Lay et al.. 2007). Upregulation of pro-inflammatory cytokines was also demonstrated in
32 sputum of O3-exposed subjects (Alexis etal.. 2010). One of these cytokines, IL-12p70,
33 correlated with numbers of dendritic-like cells in the sputum, and is an indicator of
34 dendritic cell activation (Alexis et al.. 2010). These authors have previously reported that
35 exposure of exercising human subjects to 400 ppb O3 for 2 hours resulted in activation of
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1 monocytes and macrophages (Lav et al.. 2007). which could play a role in exacerbating
2 existing asthma by activating allergen-specific memory T cells. The current study
3 confirms these findings and extends them by suggesting a potential mechanism whereby
4 O 3-activated dendritic cells could stimulate naive T-cells to promote the development of
5 asthma (Alexis et al.. 2010). A companion study by these same investigators (described in
6 detail in Section 5.4.2.1) provides evidence of dendritic cell activation, measured as
7 increased expression of HLA-DR, in a subset of the human subjects (GSTM1 null)
8 exposed to 400 ppb O3 for 2 hours with intermittent exercise (Alexis et al.. 2009). Since
9 dendritic cells are a link between innate and adaptive immunity, these studies provide
10 evidence for an O3-mediated interaction between the innate and adaptive immune
11 systems.
12 Another recent study linked O3-mediated activation of the innate immune system to the
13 development of non-specific AHR in a mouse model (Pichavant et al.. 2008). Repeated
14 exposure to 1 ppm O3 for 3 hours (3 days over a 5 day period) induced non-specific AHR
15 measured 24 hours following the last exposure (Pichavant et al.. 2008). This response
16 was found to require NKT cells, which are effector lymphocytes of innate immunity, as
17 well as IL-17 and airways neutrophilia (Pichavant et al.. 2008). Since glycolipids such as
18 galactosyl ceramide are ligands for the invariant CD 1 receptor on NKT cells and serve as
19 endogenous activators of NKT cells, a role for O3-oxidized lipids in activating NKT cells
20 was proposed (Pichavant et al.. 2008). The authors contrasted this innate immunity
21 pathway with that of allergen-provoked specific AHR which involves adaptive immunity,
22 the cytokines IL-4, IL -13, IL-17, and airways eosinophilia (Tichavant et al.. 2008).
23 Interestingly, NKT cells were required for both the specific AHR provoked by allergen
24 and the non-specific AHR provoked by O3 (Tichavant et al.. 2008). Different cytokine
25 profiles of the NKT cells from allergen and O3-exposed mice was proposed to account
26 for the different pathways (Tichavant et al.. 2008). More recently, NKT cells have been
27 found to function in both innate and adaptive immunity (Vivier et al.. 2011).
28 An interaction between allergen and O3 in the induction of nonspecific AHR was shown
29 in another animal study (Larsen et al.. 2010). Mice were sensitized with the aerosolized
30 allergen OVA on 10 consecutive days followed by exposure to O3 (0.1-0.5 ppm for 3
31 hours) (Larsen etal.. 2010). While allergen sensitization alone did not alter airways
32 responsiveness to a nonspecific challenge, O3 exposure of sensitized mice resulted in
33 nonspecific AHR at 6- and 24-hours postexposure (Larsen et al.. 2010). The effects of O3
34 on AHR were independent of airways eosinophilia and neutrophilia (Larsen et al.. 2010).
3 5 However, OVA pretreatment led to goblet cell metaplasia which was enhanced by O3
36 exposure (Larsen et al.. 2010). It should be noted that OVA sensitization using only
37 aerosolized antigen in this study is less common than the usual procedure for OVA
3 8 sensitization achieved by one or more initial systemic injections of OVA and adjuvant
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1 followed by repeated inhalation exposure to OVA. This study also points to an interaction
2 between innate and adaptive immune systems in the development of the AHR response.
3 Furthermore, O3 was found to act as an adjuvant for allergic sensitization (Hollingsworth
4 etal.. 2010). Oropharyngeal aspiration of OVA on day 0 and day 6 failed to lead to
5 allergic sensitization unless mice were first exposed to 1 ppm O3 for 2 hours
6 (Hollingsworth et al.. 2010). The O3-mediated response involved Th2 (IL-4, IL-5 and IL-
7 9) and Thl7 cytokines (IL-17) and was dependent on a functioning TLR4 (Hollingsworth
8 etal.. 2010). Ozone exposure also activated OVA-bearing dendritic cells in the thoracic
9 lymph nodes, as measured by the presence of the CD86 surface marker, which suggests
10 naive T cell stimulation and the involvement of Th2 pathways (Hollingsworth et al..
11 2010). Thus the adjuvant effects of O3 may be due to activation of both innate and
12 adaptive immunity.
13 Priming of the innate immune system by O3 was reported by Hollingsworth et al. (2007).
14 In this study, exposure of mice to 2 ppm O3 for 3 hours led to nonspecific AHR at 24-
15 and 48-hours postexposure, an effect which subsided by 72 hours (Hollingsworth et al.,
16 2007). However, in mice treated with aerosolized endotoxin immediately following O3
17 exposure, AHR was greatly enhanced at 48-and 72-hours postexposure (Hollingsworth et
18 al., 2007). In addition, O3 pre-exposure was found to reduce the number of inflammatory
19 cells in the BALF, to increase cytokine production and total protein in the BALF and to
20 increase systemic IL-6 following exposure to endotoxin (Hollingsworth et al., 2007).
21 Furthermore, O3 stimulated the apoptosis of alveolar macrophages 24-hours
22 postexposure, an effect which was greatly enhanced by endotoxin treatment. Apoptosis of
23 circulating blood monocytes was also observed in response to the combined exposures
24 (Hollingsworth et al., 2007). Ozone pre-exposure enhanced the response of lung
25 macrophages to endotoxin (Hollingsworth et al.. 2007). Taken together, these findings
26 demonstrated that O3 exposure increased innate immune responsiveness to endotoxin.
27 The authors attributed these effects to the increased surface expression of TLR4 and
28 increased signaling in macrophages observed in the study (Hollingsworth et al., 2007). It
29 was proposed that the resulting decrease in airway inflammatory cells could account for
30 O3-mediated decreased clearance of bacterial pathogens observed in numerous animal
31 models (Hollingsworth et al.. 2007).
32 More recently, these authors demonstrated that hyaluronan contributed to the O3-primed
33 response to endotoxin (Li etal., 2010). In this study, exposure of mice to 1 ppm O3 for 3
34 hours resulted in enhanced responses to endotoxin, which was mimicked by intratracheal
35 instillation of hyaluronan fragments (Li etal.. 2010). Hyaluronan, like O3, was also found
36 to induce TLR4 receptor peripheralization in the macrophage membrane (Li etal.. 2010;
37 Hollingsworth et al., 2007). an effect which is associated with enhanced responses to
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1 endotoxin. This study and previous ones by the same investigators showed elevation of
2 BALF hyaluronan in response to O3 exposure (Garantziotis et al.. 2010; Li etal.. 2010;
3 Garantziotis et al.. 2009). providing evidence that the effects of O3 on innate immunity
4 are at least in part mediated by hyaluronan fragments. The authors note that excessive
5 TLR4 signaling can lead to lung injury and suggest that O3 may be responsible for an
6 exaggerated innate immune response which may underlie lung injury and decreased host
7 defense (Li etal.. 2010).
8 Activation or upregulation of the immune system has not been reported in all studies.
9 Impaired antigen-specific immunity was demonstrated following subacute O3 exposure
10 (0.6 ppm, 10 h/day for 15 days) in mice (Feng et al.. 2006). Specifically, O3 exposure
11 altered the lymphocyte subset and cytokine profile and impacted thymocyte early
12 development leading to immune dysfunction. Further, recent studies demonstrated SP-A
13 oxidation in mice exposed for 3-6 hours to 2 ppm O3. SP-A is an important innate
14 immune protein which plays a number of roles in host defense including acting as
15 opsonin for the recognition of some pathogens (Hague etal.. 2009). These investigations
16 found that O 3 -mediated carbonylation of SP-A was associated with impaired macrophage
17 phagocytosis in vitro (Mikerov et al.. 2008b). Furthermore, O3 exposure (2 ppm for 3
18 hours) in mice was found to increase susceptibility to pneumonia infection in mice
19 through an impairment of SP-A dependent phagocytosis (Mikerov et al.. 2008a; Mikerov
20 etal.. 2008c).
21 Taken together, results of recent studies provide evidence that O3 alters host
22 immunologic response and leads to immune system dysfunction through its effects on
23 innate and adaptive immunity.
5.3.7 Airways remodeling
24 As noted above, the degree of airways inflammation due to O3 may have important long-
25 term consequences since airways inflammation is often accompanied by tissue injury
26 (Bahnes et al.. 1996). The nasal airways, conducting airways and distal airways (i.e.
27 respiratory bronchioles or centriacinar region depending on the species) have all been
28 identified as sites of O3-mediated injury and inflammation (Mudway and Kelly. 2000). At
29 all levels of the respiratory tract, loss of sensitive epithelial cells, degranulation of
30 secretory cells, proliferation of resistant epithelial cells and neutrophilic influx have been
31 observed as a result of O3 exposure (Mudway and Kelly. 2000; Cho etal.. 1999). An
32 important study (Plopper et al.. 1998) conducted in adult rhesus monkeys (0.4 and
33 1.0 ppm O3 for 2 hours) found that 1 ppm O3 resulted in the greatest epithelial injury in
34 the respiratory bronchioles immediately postexposure although injury was observed at all
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1 of the RT sites studied except for the lung parenchyma. Exposure to 0.4 ppm O3 resulted
2 in epithelial injury only in the respiratory bronchioles.
3 Persistent inflammation and injury, observed in animal models of chronic and intermittent
4 exposure to O3, are associated with airways remodeling, including mucous cell
5 metaplasia of nasal transitional epithelium (Harkema et al.. 1999; Hotchkiss et al.. 1991)
6 and bronchiolar metaplasia of alveolar ducts (Mudway and Kelly. 2000). Fibrotic changes
7 such as deposition of collagen in the airways and sustained lung function decrements
8 especially in small airways have also been demonstrated as a response to chronic O3
9 exposure (Mudway and Kelly. 2000; Chang et al.. 1992). These effects, described in
10 detail in Section 7.2.3.1, have been demonstrated in rats exposed to levels of O3 as low as
11 0.25 ppm. Mechanisms responsible for the resolution of inflammation and the repair of
12 injury remain to be clarified and there is only a limited understanding of the biological
13 processes underlying long-term morphological changes. However, a recent study in mice
14 demonstrated a key role for the TGF-|3 signaling pathway in the deposition of collagen in
15 the airways wall following chronic intermittent exposure to 0.5 ppm O3 (Katre et al..
16 2011).
17 It should be noted that repeated exposure to O3 results in attenuation of some O3 -
18 induced responses, including those associated with the activation of neural reflexes (e.g.
19 decrements in pulmonary function), as discussed in Section 5.3.2. However, numerous
20 studies demonstrate that some markers of injury and inflammation remain increased
21 during multi-day exposures to O3. Mechanisms responsible for attenuation, or the lack
22 thereof, are incompletely understood.
5.3.8 Systemic inflammation and oxidative/nitrosative stress
23 Extrapulmonary effects of O3 have been noted for decades (U.S. EPA. 2006b). It has
24 been proposed that lipid oxidation products resulting from reaction of O3 with lipids
25 and/or cellular membranes in the ELF are responsible for systemic effects, however it is
26 not known whether they gain access to the vascular space (Chuang et al.. 2009).
27 Alternatively, extrapulmonary release of diffusible mediators may initiate or propagate
28 inflammatory responses in the vascular or in systemic compartments (Cole and Freeman.
29 2009). A role for O3 in modulating endothelin, a potent vasoconstrictor, has also been
30 proposed. Studies in rats found that exposure to 0.4 and 0.8 ppm O3 induced endothelin
31 system genes in the lung and increased circulating levels of endothelin (Thomson et al..
32 2006; Thomson et al.. 2005). Systemic oxidative stress is suggested by studies in humans
33 which reported associations between O3 exposure and levels of plasma 8-isoprostanes
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1 and the presence of peripheral blood lymphocyte micronuclei (Chen et al.. 2007; Chen et
2 al.. 2006a).
3 Ozone-induced perturbations of the cardiovascular system were recently investigated in
4 young mice and monkeys (Chuang et al., 2009) and in rats (Kodavanti et al.. 2011;
5 Perepu et al.. 2010) (see Sections 6.3.3.2 and 7.3.1.2). These are the first studies to
6 suggest that systemic oxidative stress and inflammation play a mechanistic role in O3-
7 induced effects on the systemic vascular and heart. Exposure to 0.5 ppm O3 for 5 days
8 resulted in oxidative/nitrosative stress, vascular dysfunction and mitochondrial DNA
9 damage in the aorta (Chuang et al.. 2009). Chronic exposure to 0.8 ppm O3 resulted in an
10 enhancement of inflammation and lipid peroxidation in the heart following an ischemia-
11 reperfusion challenge (Perepu etal.. 2010). In addition, chronic intermittent exposure to
12 0.4 ppm O3 increased aortic levels of mRNA for biomarkers of oxidative stress,
13 thrombosis, vasoconstriction and proteolysis and aortic lectin-like oxidized4ow density
14 lipoprotein receptor-l(LOX-l) mRNA and protein levels (Kodavanti et al., 2011). The
15 latter study suggests a role for circulating oxidized lipids in mediating the effects of O3.
16 Systemic inflammation and oxidative/nitrosative stress may similarly affect other organ
17 systems as well as the plasma compartment. Circulating cytokines have the potential to
18 enter the brain through diffusion and active transport and to contribute to
19 neuroinflammation, neurotoxicity, cerbrovascular damage and a break-down of the blood
20 brain barrier (Block and Calderon-Garciduenas. 2009) (see Sections 6.4 and 7.5). They
21 can also activate neuronal afferents (Block and Calderon-Garciduenas. 2009). Vagal
22 afferent pathways originating in the respiratory tract may also be responsible for O3-
23 mediated activation of nucleus tractus solitarius neurons which resulted in neuronal
24 activation in stress-responsive regions of the CNS in rats (0.5 or 2 ppm O3 for 1.5-120
25 hours) (Gackiere et al.. 2011). Recent studies have demonstrated O3-induced brain lipid
26 peroxidation, cytokine production in the brain and upregulated expression of VEGF in
27 rats (0.5 ppm O3, 3 hours or 0.25-0.5 ppm O3, 4 h/day, 15-60 days) (Guevara-Guzman et
28 al.. 2009; Araneda et al., 2008; Perevra-Munoz et al., 2006). Further, O3-induced
29 oxidative stress resulted in increased plasma lipid peroxides (0.25 ppm, 4h/day, 15-60
30 days) (Santiago-Lopez et al., 2010). which was correlated with damage to specific brain
31 regions (Pereyra-Munoz et al.. 2006).
32 Oxidative stress is one mechanism by which testicular and sperm function is disrupted
33 (see Section 7.4.1). Oxidative stress may inhibit testicular steroidogenesis leading to
34 decreased testosterone levels (Diemer et al.. 2003). It may decrease sperm quality by lipid
3 5 peroxidation of sperm plasma membrane which leads to impaired sperm mobility
36 (Agarwal et al.. 2003). Further, it may damage DNA in the sperm nucleus leading to
37 apoptosis and a decline in sperm counts (Agarwal et al.. 2003). Since oxidative stress is a
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1 key event underlying many of the health effects of O3, it is possible that sperm quality
2 and quantity may be impacted by this mechanism (Sokol et al.. 2006).
3 A role for plasma antioxidants in modulating O3-induced respiratory effects was
4 suggested by a recent study (Aibo etal.. 2010). In this study, pretreatment of rats with a
5 high dose of acetaminophen resulted in increased levels of plasma cytokines and the
6 influx of inflammatory cells into the lung following O3 exposure (0.25-0.5 ppm, 6 hours)
7 (Aiboet al.. 2010). These effects were not observed in response to O3 alone.
8 Furthermore, acetaminophen-induced liver injury was exacerbated by O3 exposure. A
9 greater increase in hepatic neutrophil accumulation and greater alteration in gene
10 expression profiles was observed in mice exposed to O3 and acetaminophen compared
11 with either exposure alone (Aibo et al.. 2010). Although not measured in this study,
12 glutathione depletion in the liver is known to occur in acetaminophen toxicity. Since liver
13 glutathione is the source of plasma glutathione, acetaminophen treatment may have
14 lowered plasma glutathione levels and altered the redox balance in the vascular
15 compartment. These findings indicate interdependence between respiratory tract, plasma
16 and liver responses to O3, possibly related to glutathione status.
5.3.9 Impaired alveolar-arterial 02 transfer
17 O3 may impair alveolar-arterial oxygen transfer and reduce the supply of arterial oxygen
18 to the myocardium. This may have a greater impact in individuals with compromised
19 cardiopulmonary systems. Gong et al. (1998) provided evidence of a small decrease in
20 arterial oxygen saturation in human subjects exposed for 3 hours to 300 ppb O3 while
21 exercising. In addition, Delaunois et al. (1998) demonstrated pulmonary vasoconstriction
22 in O3-exposed rabbits (0.4 ppm, 4 hours). Although of interest, the contribution of this
23 pathway to O3-induced cardiovascular effects remains uncertain.
5.3.10 Summary
24 This section summarizes the modes of action and toxicity pathways resulting from O3
25 inhalation (Figure 5-9). These pathways provide a mechanistic basis for the health effects
26 which are described in detail in Chapters 6 and 7. Three distinct short-term responses
27 have been well-characterized in humans challenged with O3: decreased pulmonary
28 function, airways inflammation, and increased bronchial reactivity. In addition, O3
29 exposure exacerbates, and possibly also causes, asthma and allergic airways disease in
30 humans. Animal studies have demonstrated airways remodeling and fibrosis in response
31 to chronic and intermittent O3 exposures and a wide range of other responses. While the
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1
2
respiratory tract is the primary target tissue, cardiovascular and other organ effects occur
following short- and long-term exposures of animals to O3.
Ozone + Respiratory Tract
Formation of secondary oxidation products
Activation
of neural
reflexes
Initiation of
inflammation
Sensitization
of bronchial
smooth muscle
Systemic inflammation and
oxidative/nitrosative stress
Extrapulmonary Effects
Decrements in pulmonary function
Pulmonary inflammation/oxidative stress
Increases in airways permeability
Airways hyperresponsiveness
Exacerbation/induction of asthma
Decreased host defenses
Epithelial metaplasia and fibrotic airways
Altered lung development
Figure 5-9 The modes of action/possible pathways underlying the health
effects resulting from inhalation exposure to O3.
4
5
6
7
10
11
12
13
14
The initial key event in the toxicity pathway of O3 is the formation of secondary
oxidation products in the respiratory tract. This involves direct reactions with components
of the ELF and/or plasma membranes of cells residing in the respiratory tract. The
resulting secondary oxidation products transmit signals to the epithelium, nociceptive
sensory nerve fibers and, if present, dendritic cells, mast cells and eosinophils. Thus, O3
effects are mediated by components of ELF and by the multiple cell types found in the
respiratory tract. Further, oxidative stress is an implicit part of this initial key event.
Another key event in the toxicity pathway of O3 is the activation of neural reflexes which
lead to decrements in pulmonary function (see Section 6.2.1). Evidence is accumulating
that secondary oxidation products are responsible for this effect. Eicosanoids have been
implicated in humans, while both eicosanoids and aldehydes are effective in animal
models. Different receptors on bronchial C-fibers have been shown to mediate separate
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September 2011
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1 effects of O3 on pulmonary function. Nociceptive sensory nerves are involved in the
2 involuntary truncation of respiration which results in decreases in FVC, FEVi, tidal
3 volume and pain upon deep inspiration. Opioids block these responses while atropine has
4 only a minimal effect. New evidence in an animal model suggests that TRPA1 receptors
5 on bronchial C-fibers mediate this pathway. Ozone exposure also results in activation of
6 vagal sensory nerves and a mild increase in airway obstruction measured as increased
7 sRaw. Atropine and (3-adrenergic agonists greatly inhibit this response in humans
8 indicating that the airway obstruction is due to bronchoconstriction. Other studies in
9 humans implicated SP release from bronchial C-fibers resulting in airway narrowing due
10 to either neurogenic edema or bronchoconstriction. New evidence in an animal model
11 suggests that the SP-NK receptor pathway caused bronchoconstriction following O3
12 exposure.
13 Initiation of inflammation is also a key event in the toxicity pathway of O3. Secondary
14 oxidation products, as well as chemokines and cytokines elaborated by airway epithelial
15 cells and macrophages, have been implicated in the initiation of inflammation. Vascular
16 endothelial adhesion molecules may also play a role. Work from several laboratories in
17 using human subjects and animal models suggest that O3 triggers the release of
18 tachykinins such as SP from airway sensory nerves which could contribute to
19 downstream effects including inflammation (see Sections 6.2.3 and 7.2.4). Airways
20 neutrophilia has been demonstrated in BALF, mucosal biopsy and induced sputum
21 samples. Influx of mast cells, monocytes and macrophages also occur. Inflammation
22 further contributes to O3-mediated oxidative stress. Recent investigations show that O3
23 exposure leads to the generation of hyaluronan fragments from high molecular weight
24 polymers of hyaluronan normally found in the ELF in mice. Hyaluronan activates TLR4
25 and CD44-dependent signaling pathways in macrophages, and results in an increased
26 number of macrophages in the BALF. Activation of these pathways occurs later than the
27 acute neutrophilic response suggesting that they may contribute to longer-term effects of
28 O3. The mechanisms involved in clearing O3-provoked inflammation remain to be
29 clarified. It should be noted that inflammation, as measured by airways neutrophilia, is
30 not correlated with decrements in pulmonary function as measured by spirometry.
31 A fourth key event in the toxicity pathway of O3 is alteration of epithelial barrier
32 function. Increased permeability occurs as a result of damage to tight junctions between
33 epithelial cells subsequent to O3-induced injury and inflammation. It may play a role in
34 allergic sensitization and in AHR (see Sections 6.2.2, 6.2.6, and 7.2.5). Tachykinins
35 mediate this response while antioxidants confer protection. Genetic susceptibility has
36 been associated with a functioning TLR4 gene and with iNOS.
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1 A fifth key event in the toxicity pathway of O3 is the sensitization of bronchial smooth
2 muscle.
3 Increased bronchial reactivity can be both a rapidly occurring and a persistent response.
4 The mechanisms responsible for early and later AHR are not well-understood (see
5 Section 6.2.2). One proposed mechanism of sensitization, O3-induced increases in
6 epithelial permeability, would improve access of agonist to smooth muscle receptors. The
7 evidence for this mechanism is not consistent. Another proposed mechanism, for which
8 there is greater evidence, is neurally-mediated sensitization. In humans exposed to O3,
9 atropine blocked the early AHR response indicating the involvement of cholinergic
10 postganglionic pathways. Animal studies demonstrated that O3-induced AHR involved
11 vagally-mediated responses and local axon reflex responses through bronchopulmonary
12 C-fiber-mediated release of SP. Later phases of increased bronchial reactivity may
13 involve the induction of IL-1(3 which in turn upregulates SP production. In guinea pigs,
14 eosinophil-derived major basic protein contributed to the stimulation of cholinergic
15 postganglionic pathways. A novel role for hyaluronan in mediating the later phase effects
16 O 3 -induced AHR has recently been demonstrated. Hyaluronan fragments stimulated AHR
17 in a TLR4- and CD44 receptor-dependent manner. Tachykinins and secondary oxidation
18 products of O3 have been proposed as mediators of the early response and inflammation-
19 derived products have been proposed as mediators of the later response. Inhibition of
20 arachidonic acid metabolism was ineffective in blocking O3-induced AHR in humans
21 while in animal models mixed results were found. Other cytokines and chemokines have
22 been implicated in the AHR response to O3 in animal models.
23 A sixth key event in the toxicity pathway of O3 is the modification of innate/adaptive
24 immunity. While the majority of evidence for this key event comes from animal studies,
25 there are several studies suggesting that this pathway may also be relevant in humans. O3
26 exposure of human subjects resulted in recruitment of activated innate immune cells to
27 the airways. This included macrophages and monocytes with increased expression of cell
28 surface markers characteristic of innate immunity and antigen presentation, the latter of
29 which could contribute to exaggerated T cell responses and the promotion of an allergic
30 phenotype. Evidence of dendritic cell activation was observed in GSTM1 null human
31 subjects exposed to O3, suggesting O3-mediated interaction between the innate and
32 immune systems. Animal studies further linked O3-mediated activation of the innate
33 immune system to the development of nonspecific AHR, demonstrated an interaction
34 between allergen and O3 in the induction of nonspecific AHR, and found that O3 acted as
35 an adjuvant for allergic sensitization through the activation of both innate and adaptive
36 immunity. Priming of the innate immune system by O3 was reported in mice. This
37 resulted in an exaggerated response to endotoxin which included enhanced TLR4
38 signaling in macrophages. Ozone-mediated impairment of the function of SP-A, an innate
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1 immune protein, has also been demonstrated. Taken together these studies provide
2 evidence that O3 can alter host immunologic response and lead to immune system
3 dysfunction. These mechanisms may underlie the exacerbation and induction of asthma
4 (see Sections 6.2.6 and 7.2.1), as well as decreases in host defense (see Sections 6.2.5 and
5 7.2.6).
6 Another key event in the toxicity pathway of O3 is airways remodeling. Persistent
7 inflammation and injury, which are observed in animal models of chronic and
8 intermittent exposure to O3, are associated with morphologic changes such as mucous
9 cell metaplasia of nasal epithelium, bronchiolar metaplasia of alveolar ducts and fibrotic
10 changes in small airways (see Section 7.2.3). Mechanisms responsible for these responses
11 are not well-understood. However a recent study in mice demonstrated a key role for the
12 TGF-(3 signaling pathway in the deposition of collagen in the airway wall following
13 chronic intermittent exposure to O3.
14 Systemic inflammation and vascular oxidative/nitrosative stress are also key events in the
15 toxicity pathway of O3. Extrapulmonary effects of O3 occur in numerous organ systems,
16 including the cardiovascular, central nervous, reproductive and hepatic systems (see
17 Sections 6.3 to 6.5 and 7.3 to 7.5). It has been proposed that lipid oxidation products
18 resulting from reaction of O3 with lipids and/or cellular membranes in the ELF are
19 responsible for systemic responses, however it is not known whether they gain access to
20 the vascular space. Alternatively, release of diffusible mediators from the lung into the
21 circulation may initiate or propagate inflammatory responses in the vascular or in
22 systemic compartments. Systemic oxidative stress is suggested by studies in humans
23 which reported associations between O3 exposure and levels of plasma 8-isoprostanes
24 and the presence of peripheral blood lymphocyte micronuclei.
5.4 Interindividual Variability in Response
25 Responses to O3 exposure are variable within the population and the basis for this
26 variability is not clear (Mudway and Kelly. 2000). Both dosimetric and mechanistic
27 factors are likely to contribute to this variability and are discussed below.
5.4.1 Dosimetric Considerations
28 Two studies have investigated the correlation of O3 uptake with the pulmonary function
29 responses to O3 exposure (Reeser et al., 2005; Gerrityetal.. 1994). These studies found
30 that the large subject-to-subject variability in %AFEVi response to O3 does not appear to
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1 have a dosimetric explanation. Reeser et al. (2005) found no significant relationship
2 between %AFEVi and fractional absorption of O3 using the bolus method. Contrary to
3 previous findings, the percent change in dead space volume of the respiratory tract
4 (%AVD) did not correlate with O3 uptake, possibly due to the contraction of dead space
5 caused by airway closure. Gerrity et al. (1994) found that intersubject variability in FEVi
6 and airway resistance was not related to differences in the O3 dose delivered to the lower
7 airways, whereas minute ventilation was predictive of FEVi decrement. No study has yet
8 demonstrated that subjects show a consistent pattern of O3 retention when re-exposed
9 over weeks of time, as has been shown to be the case for the FEVi response, or that
10 within-subject variation in FEVi response is related to fluctuations in O3 uptake.
11 A delay in onset of O3-induced pulmonary function responses has been noted in
12 numerous studies. Recently the delay was characterized in terms of changes in fB
13 (Schelegle et al.. 2007). In humans exposed for 1-2 hours to 120-350 ppb O3 while
14 exercising, no change in fB was observed until a certain cumulative inhaled dose of O3
15 had been reached. Subsequently, the magnitude of the change in fB was correlated with
16 the inhaled dose rate (Schelegle et al.. 2007). These investigators proposed that initial
17 reactions of O3 with ELF resulted in a time-dependent depletion of ELF antioxidants, and
18 that activation of neural reflexes occurred only after the antioxidant defenses were
19 overwhelmed (Schelegle et al.. 2007).
20 Other studies investigated the relationship between O3 dose and cellular injury. In two
21 studies, the initial cellular injury was found to correlate with the site-specific O3 dose.
22 Contained within the CAR, the respiratory bronchioles were confirmed as the site
23 receiving the greatest O3 dose (18O mass/lung weight) and sustained the greatest initial
24 cellular injury in O3 (0.4 and 1.0 ppm for 2 hours) exposed resting rhesus monkeys
25 (Plopper et al.. 1998). The respiratory bronchioles, having the highest concentration of
26 local O3 dose, were also the site of significant GSH reduction. In addition, a study in
27 isolated perfused rat lungs found greater injury in conducting airways downstream of
28 bifurcations where local doses of O3 were higher (Postlethwait et al.. 2000).
29 Further, the degree of inflammation in rats has been correlated with 18O-labeled O3 dose
30 markers in the lower lung. In female rats exposed to 0.8 ppm O3 for 4 hours, BAL
31 neutrophil number and 18O reaction product were directly proportional (Gunnison and
32 Hatch. 1999). Kari et al. (1997) observed that a 3-week caloric restriction (75%) in rats
33 abrogated the toxicity of O3 (2 ppm, 2 hours), measured as BALF increases in protein,
34 fibronectin and neutrophils, which was seen in normally fed rats. Accompanying this
35 resistance to O3 toxicity was a reduction (30%) in the accumulation of 18O reaction
36 product in the lungs. These investigations also demonstrated an inverse relationship
37 between AH2 levels and O3 dose and provided evidence for AH2 playing a protective
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1 role following O3 exposure in these studies. Pregnant and lactating rats had lower AH2
2 content in BALF and exhibited a greater increase in accumulation of 18O reaction
3 products compared with pre-pregnant rats in response to O3 exposure (Gunnison and
4 Hatch. 1999). In the calorie restricted model, a 30% higher basal BALF AH2
5 concentration and a rapid accumulation of AH2 into the lungs to levels 60% above
6 normal occurred as result of O3 exposure (Kari etal.. 1997). However, this relationship
7 between AH2 levels and O3 dose did not hold up in every study. Aging rats (9 and 24
8 months old) had 49% and 64% lower AH2 in lung tissue compared with month-old rats
9 but the aging-induced AH2 loss did not increase the accumulation of 18O reaction
10 products following O3 exposure (0.4-0.8 ppm, 2-6 hours) (Vincent et al.. 1996a).
11 Interindividual variability in the neutrophilic response has been noted in human subjects
12 (Holz etal.. 1999; Devlin etal.. 1991; Schelegle et al.. 1991). One study demonstrated a
13 threefold difference in airways neutrophilia, measured as percent of total cells in
14 proximal BALF, among human subjects exposed to 300 ppb O3 for 1 hour while
15 exercising (Schelegle et al.. 1991). Another study reported a 20-fold difference in BAL
16 neutrophils following exposure to 80-100 ppb O3 for 6.6 hours in exercising human
17 subjects (Devlin et al.. 1991). Reproducibility of intra-individual responses to 1-hour
18 exposure to 250 ppb O3, measured as sputum neutrophilia, was demonstrated by Holz
19 (1999). Few studies have examined the dose- or concentration-responsiveness of airways
20 neutrophilia in O3-exposed humans (Holz et al.. 1999; Devlin etal.. 1991). No
21 concentration-responsiveness was observed in healthy human subjects exposed for 1 hour
22 to 125-250 ppb O3 and a statistically significant increase in sputum neutrophilia was
23 observed only at the higher dose (Holz etal.. 1999). However, concentration-dependent
24 and statistically significant increases in BAL neutrophils and the inflammatory mediator
25 IL-6 were reported following exposure to 80 and 100 ppb O3 for 6.6 hours in exercising
26 humans (Devlin et al.. 1991). Additional evidence is provided by a meta-analysis of the
27 O3 dose-inflammatory response in controlled human exposure studies involving exposure
28 to 80-600 ppb O3 for 60-396 minutes (Mudwav and Kelly. 2004b). Results demonstrated
29 a linear relationship between inhaled O3 dose (determined as the product of
30 concentration, ventilation and time) and BAL neutrophils at 0-6 hours and 18-24 hours
31 following O3 exposure (Mudway and Kelly. 2004b).
32 Collectively these studies demonstrate a correlation between dose and response for some
33 O3-induced effects and suggest a role for ELF antioxidants in modulating the dose to
34 tissue. The lack of correlation between O3-induced effects and calculated O3 dose may be
35 a result of interindividual differences in TB volume.
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5.4.2 Mechanistic Considerations
1 There was a large range of pulmonary function responses to O3 among healthy young
2 adults exposed for 4 hours to 200 ppb O3 or for 1.5 hours to 420 ppb O3 while exercising
3 (Hazucha et al.. 2003; Balmes et al.. 1996). Since individual responses were relatively
4 consistent across time, it was thought that responsiveness reflected an intrinsic
5 characteristic of the subject (Mudway and Kelly. 2000). Older adults were generally not
6 responsive to O3 (Hazucha et al.. 2003). while obese young women may have been more
7 responsive than lean young women (420 ppb, 1.5 hours, while exercising) (Bennett et al..
8 2007). The lack of spirometric responsiveness was not attributable to the presence of
9 endogenous endorphins, which could potentially block C-fiber stimulation by O3, as
10 demonstrated in a study involving intravenous administration of naloxone immediately
11 following the O3 exposure (420 ppb, 2 hours, while exercising) to weak responders
12 (Passannante et al.. 1998). Inflammation and other responses to O3 were also
13 characterized by a large degree of interindividual variability. Currently, the mechanisms
14 underlying this variability are not known. It has been proposed that some of the variation
15 in responses may be genetically determined (Yang et al.. 2005a). The role of gene-
16 environment interactions, pre-existing diseases and conditions, nutritional status,
17 lifestage, attenuation, and co-exposures in modulating responses to O3 are discussed
18 below.
5.4.2.1 Gene-Environment Interactions
19 The significant interindividual variation in responses to O3 infers that genetic background
20 is an important determinant of susceptibility to O3 (Cho and Kleeberger. 2007;
21 Kleeberger et al.. 1997) (see also Section 8.4). Strains of mice which are prone or
22 resistant to O3-induced effects have been used to systematically identify candidate
23 susceptibility genes. Genome wide linkage analyses (also known as positional cloning)
24 demonstrated quantitative trait loci for O3-induced lung inflammation and
25 hyperpermeability on chromosome 17 (Kleeberger et al.. 1997) and chromosome 4
26 (Kleeberger et al.. 2000). respectively, using these recombinant inbred strains of mice and
27 exposures to 0.3 ppm O3 for up to 72 hours. More specifically, these studies found that
28 Tnf, whose protein product is the inflammatory cytokine TNF-a, and Tlr4, whose protein
29 product is TLR4, were candidate susceptibility genes (Kleeberger et al.. 2000; Kleeberger
30 etal.. 1997). Other studies, which used targeted deletion, identified genes encoding iNOS
31 and heat shock proteins as TLR4 effector genes (Bauer etal.. 2011; Kleeberger et al..
32 2001) and found that IL-10 protects against O3-induced pulmonary inflammation
33 (Backus etal.. 2010). Investigations in inbred mouse strains found that differences in
34 expression of certain proteins, such as CCSP (1.8 ppm O3 for 3 hours) (Broeckaert et al..
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1 2003) and MARCO (0.3 ppm O3 for up to 48 hours) (Dahl et al.. 2007). were responsible
2 for phenotypic characteristics, such as epithelial permeability and scavenging of oxidized
3 lipids, respectively, which confer sensitivity to O3.
4 Genetic polymorphisms have received increasing attention as modulators of O3-mediated
5 effects. Functionally relevant polymorphisms in candidate susceptibility genes have been
6 studied at the individual and population level in humans, and also in animal models.
7 Genes whose protein products are involved in antioxidant defense/oxidative stress and
8 xenobiotic metabolism, such as glutathione-S-transferase Ml (GSTM1) and
9 NADPH:quinone oxidoreductase 1 (NQO1), have also been a major focuses of these
10 efforts. This is because oxidative stress resulting from O3 exposure is thought to
11 contribute to the pathogenesis of asthma, and because xenobiotic metabolism detoxifies
12 secondary oxidation products formed by O3 which contribute to oxidative stress (Islam et
13 al.. 2008). TNF-a is of interest since it is linked to a candidate O3 susceptibility gene and
14 since it plays a key role in initiating airways inflammation (Li et al.. 2006d).
15 Polymorphisms of genes coding for GSTM1, NQO1 and TNF-a have been associated
16 with altered susceptibility to O3-mediated effects (Li et al.. 2006d; Yang et al.. 2005a;
17 Romieu et al.. 2004a: Corradi et al.. 2002; Bergamaschi et al.. 2001). Additional studies
18 have focused on functional variants in other genes involved in antioxidant defense such
19 as catalase (CAT), myeloperoxidase, heme oxygenase (HMOX-1) and manganese
20 superoxide dismutase (MnSOD) (Wenten et al.. 2009; Islam et al.. 2008). These studies
21 are discussed below.
22 GSTM1 is a phase II antioxidant enzyme which is transcriptionally regulated by NF-E2-
23 related factor 2-antioxidant response element (Nrf2-ARE) pathway. A large proportion
24 (40-50%) of the general public (across ethnic populations) has the GSTMl-null genotype,
25 which has been linked to an increased risk of health effects due to exposure to air
26 pollutants (London. 2007). A role for GSTs in metabolizing electrophiles such as 4-
27 hydroxynonenal, which is a secondary oxidation product formed following O3 exposure,
28 has been demonstrated (Awasthi et al.. 2004). A recent study found that the GSTM1
29 genotype modulated the time course of the neutrophilic inflammatory response following
30 acute O3 exposure (400 ppb for 2 hours with intermittent exercise) in healthy adults
31 (Alexis et al.. 2009). In GSTMl-null and -sufficient subjects, O3-induced sputum
32 neutrophilia was similar at 4 hours. However, neutrophilia resolved by 24 hours in
33 sufficient subjects but not in GSTMl-null subjects. In contrast, no differences in 24 hour
34 sputum neutrophilia were observed between GSTMl-null and -sufficient human subjects
35 exposed to 60 ppb O3 for 2 hours with intermittent exercise (KimetaL 2011). It is not
36 known whether the effect seen at the higher exposure level (Alexis et al.. 2009) was due
37 to the persistence of pro-inflammatory stimuli, impaired production of downregulators or
38 impaired neutrophil apoptosis and clearance. However, a subsequent in vitro study by
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1 these same investigators found that GSTM1 deficiency in airway epithelial cells
2 enhanced IL-8 production in response to 0.4 ppm O3 for 4 hours (Wu etal.. 2011).
3 Furthermore, NF-KB activation was required for O3-induced IL-8 production (Wu et al..
4 2011). Since IL-8 is a potent neutrophil activator and chemotaxin, this study provides
5 additional evidence for the role of GSTM1 as a modulator of inflammatory responses due
6 to O3 exposure.
7 In addition, O3 exposure increased the expression of the surface marker CD 14 in airway
8 neutrophils of GSTMl-null subjects compared with sufficient subjects (Alexis et al..
9 2009). Furthermore, differences in airway macrophages were noted between the GSTM1-
10 sufficient and -null subjects. Nnumbers of airway macrophages were decreased at 4 and
11 24 hours following O3 exposure in GSTM1-sufficient subjects (Alexis et al.. 2009).
12 Airway macrophages in GSTMl-null subjects were greater in number and found to have
13 greater oxidative burst and phagocytic capability than those of sufficient subjects. Airway
14 macrophages and dendritic cells from GSTMl-null subjects exposed to O3 expressed
15 higher levels of the surface marker HLA-DR, suggesting activation of the innate immune
16 system (Alexis et al., 2009). These differences in inflammatory responses between the
17 GSTMl-null and -sufficient subjects may provide biological plausibility for the
18 differences in O3-mediated effects reported in controlled human exposure studies
19 (Corradi et al.. 2002; Bergamaschi etal.. 2001). It should also be noted that GSTM1
20 genotype did not affect the acute pulmonary function (i.e. spirometric) response to O3
21 which provides additional evidence for separate mechanisms underlying O3's effects on
22 pulmonary function and inflammation in adults (Alexis et al.. 2009). However, GSTM1-
23 null asthmatic children were previously found to be more at risk of O3-induced effects on
24 pulmonary function than GSTM 1-sufficient asthmatic children (Romieu et al.. 2004a).
25 Another enzyme involved in the metabolism of secondary oxidation products is NQO1.
26 NQO1 catalyzes the 2-electron reduction by NADPH of quinones to hydroquinones.
27 Depending on the substrate, it is capable of both protective detoxification reactions and
28 redox cycling reactions resulting in the generation of reactive oxygen species. A recent
29 study using NQO 1 -null mice demonstrated that NQO 1 contributes to O 3 -induced
30 oxidative stress, AHR and inflammation following a 3-hour exposure to 1 ppm O3
31 rVbynow et al.. 2009). These experimental results may provide biological plausibility for
32 the increased biomarkers of oxidative stress and increased pulmonary function
33 decrements observed in O3-exposed individuals bearing both the wild-type NQO1 gene
3 4 and the null GSTM 1 gene (Corradi et al.. 2002; Bergamaschi etal.. 2001).
35 Besides enzymes, other mechanisms participate in the removal of secondary oxidation
36 products formed as a result of O3 inhalation. One involves scavenging of oxidized lipids
37 via the macrophage receptor with collagenous structure (MARCO) expressed on the cell
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1 surface of alveolar macrophages. A recent study demonstrated increased gene expression
2 of MARCO in the lungs of an O3-resistant C3H mouse strain (HeJ) but not in an O3-
3 sensitive, genetically similar strain (OuJ) (Dahl et al.. 2007). Upregulation of MARCO
4 occurred in mice exposed to 0.3 ppm O3 for 24-48 hours; inhalation exposure for 6 hours
5 at this concentration was insufficient for this response. Animals lacking the MARCO
6 receptor exhibited greater inflammation and injury, as measured by BAL neutrophils,
7 protein and isoprostanes, following exposure to 0.3 ppm O3 (Dahl et al.. 2007). MARCO
8 also protected against the inflammatory effects of oxidized surfactant lipids (Dahl et al..
9 2007). Scavenging of oxidized lipids may limit O3-induced injury since ozonized
10 cholesterol species formed in the ELF (mice, 0.5-3 ppm O3, 3 hours) (Pulfer et al.. 2005;
11 Pulfer and Murphy. 2004) stimulated apoptosis and cytotoxicity in vitro (Gao et al..
12 2009b: Sathishkumar et al.. 2009; Sathishkumar et al.. 2007a: Sathishkumar et al..
13 2007b).
14 Two studies reported relationships between TNF promoter variants and O3-induced
15 effects in humans. In one study, O3-induced change in lung function was significantly
16 lower in adult subjects with TNF promoter variants -308A/A and -308G/A compared with
17 adult subjects with the variant -308G/G (Yang et al.. 2005a). This response was
18 modulated by a specific polymorphism of LTA (Yang et al.. 2005a). a previously
19 identified candidate susceptibility gene whose protein product is lymphotoxin-a
20 (Kleeberger et al.. 1997). In the second study, an association between the TNF promoter
21 variant -308G/G and decreased risk of asthma and lifetime wheezing in children was
22 found (Li et al.. 2006d). The protective effect on wheezing was modulated by ambient O3
23 levels and by GSTM1 and GSTP1 polymorphisms. The authors suggested that the
24 TNF-308 G/G genotype may have a protective role in the development of childhood
25 asthma (Li et al.. 2006d).
26 Similarly, a promoter variant of the gene HMOX-1, consisting of a smaller number of
27 (GT)n repeats, was associated with a reduced risk for new-onset asthma in non-Hispanic
28 white children (Islam et al.. 2008). The number of (GT)n repeats in this promoter has
29 been shown to be inversely related to the inducibility of HMOX-1. A modulatory effect
30 of O3 was demonstrated since the beneficial effects of this polymorphism were seen only
31 in children living in low O3 communities (Islam et al.. 2008). This study also identified
32 an association between a polymorphism of the CAT gene and increased risk of new-onset
33 asthma in Hispanic children; however no modulation by O3 was seen (Islam et al.. 2008).
34 No association was observed in this study between a MnSOD polymorphism and asthma
35 (Mam et al.. 2008).
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1 Studies to date indicate that some variability in individual responsiveness to O3 may be
2 accounted for by functional genetic polymorphisms. Further, the effects of gene-
3 environment interactions may be different in children and adults.
5.4.2.2 Pre-existing Diseases and Conditions
4 Pre-existing diseases and conditions can alter the response to O3 exposure. For example,
5 responsiveness to O3, as measured by spirometry, is decreased in smokers and individuals
6 with COPD (U.S. EPA. 2006b). Asthma and allergic diseases are of major importance in
7 this discussion. In individuals with asthma, there is increased responsiveness to
8 bronchoconstrictor challenge. This results from a combination of structural and
9 physiological factors including increased airway inner-wall thickness, smooth muscle
10 responsiveness and mucus secretion. Although inflammation is likely to contribute, its
11 relationship to AHR is not clear (U.S. EPA. 2006b). However, some asthmatics have
12 higher baseline levels of neutrophils, lymphocytes, eosinophils and mast cells in
13 bronchial washes and bronchial biopsy tissue (Stenfors et al.. 2002). It has been proposed
14 that enhanced sensitivity to O3 is conferred by the presence of greater numbers of
15 resident airway inflammatory cells in disease states such as asthma (Mudway and Kelly.
16 2000).
17 In order to determine whether asthmatics exhibit greater responses to O3, several older
18 studies compared pulmonary function in asthmatic and non-asthmatic subjects following
19 O3 exposure. Some also probed mechanisms which could account for enhanced
20 sensitivity. While the majority focused on measurements of FEVi and FVC and found no
21 differences between the two groups following exposures of 2-4 hours to 125-250 ppb O3
22 or to a 30-minute exposure to 120-180 ppb O3 by mouthpiece while exercising (Stenfors
23 et al.. 2002; Mudwavetal.. 2001; Holz et al.. 1999; Scannell et al.. 1996; Koenig et al..
24 1987; Linn et al.. 1978). there were notable exceptions. In one study, greater airway
25 obstruction in asthmatics compared with non-asthmatic subjects was observed
26 immediately following a 2-hour exposure to 400 ppb O3 with intermittent exercise (Kreit
27 etal.. 1989). These changes were measured as statistically significant greater decreases in
28 FEVi and in FEF 25.75 (but not in FVC) in the absence of a bronchoconstrictor challenge
29 (Kreit et al.. 1989). These results suggest that this group of asthmatics responded to
30 O3-exposure with a greater degree of vagally-mediated bronchoconstriction compared
31 with the non-asthmatics. A second study demonstrated a statistically significant greater
32 decrease in FEVi and in FEVi/FVC (but not in FVC) in asthmatics compared with non-
33 asthmatics exposed to 160 ppb O3 for 7.6 hours with light exercise (Horstman et al..
34 1995). These responses were accompanied by wheezing and inhaler use in the asthmatics
35 (Horstman et al.. 1995). Aerosol bolus dispersion measurements demonstrated a
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1 statistically significant greater change in asthmatics compared with non-asthmatics,
2 which was suggestive of O3-induced small airway dysfunction (Horstman et al.. 1995).
3 Furthermore, a statistically significant correlation was observed between the degree of
4 baseline airway status and the FEVi response to O3 in the asthmatic subjects (Horstman
5 et al., 1995). A third study found similar decreases in FVC and FEVi in both asthmatics
6 and non-asthmatics exposed to 400 ppb O3 for 2 hours with mild exercise (Alexis et al..
7 2000). However, a statistically significant decrease in FEF75, a measure of small airway
8 function, was observed in asthmatics but not in non-asthmatics (Alexis et al.. 2000).
9 Taken together, these latter studies indicate that while the magnitude of restrictive type
10 spirometric decline was similar in asthmatics and non-asthmatics, that obstructive type
11 changes (i.e. bronchoconstriction) were greater in asthmatics. Further, asthmatics
12 exhibited greater sensitivity to O3 in terms of small airways function.
13 Since asthma exacerbations occur in response to allergens and/or other triggers, some
14 studies have focused on O3-induced changes in AHR following a bronchoconstrictor
15 challenge. No difference in sensitivity to methacholine bronchoprovocation was observed
16 between asthmatics and non-asthmatics exposed to 400 ppb O3 for 2 hours with moderate
17 exercise (Kreit et al.. 1989). However, increased bronchial reactivity to inhaled allergens
18 was demonstrated in mild allergic asthmatics exposed to 160 ppb for 7.6 hours, 250 ppb
19 for 3 hours and 120 ppb for 1 hour while exercising (Kehrl et al.. 1999; Torres et al..
20 1996; Molfmo et al.. 1991) and in allergen-sensitized guinea pigs following O3 exposure
21 (1 ppm, 1 hour) (Sun et al.. 1997). Similar, but modest, responses were reported for
22 individuals with allergic rhinitis (Torres et al.. 1996). Further, the contractile response of
23 isolated airways from human donor lung tissue, which were sensitized and challenged
24 with allergen, was increased by pre-exposure to 1 ppm O3 for 20 (Rouxet al.. 1999).
25 These studies provide support for O3-mediated enhancement of responses to allergens in
26 allergic subjects.
27 In terms of airways neutrophilia, larger responses were observed in asthmatics compared
28 to non-asthmatics subjects exposed to O3 in some (Balmes etal.. 1997; Scannell et al..
29 1996; Bashaetal.. 1994) but not all (Mudwavet al.. 2001) of the older studies. While
30 each of these studies involved exposure of exercising human subjects to 200 ppb O3, the
31 duration of exposure was longer (i.e. 4-6 hours) in the former studies than in the latter
32 study (2 hours). Further, statistically significantly increases in myeloperoxidase levels (an
33 indicator of neutrophil activation) in bronchial washes was observed in mild asthmatics
34 compared with non-asthmatics, despite no difference in O3-stimulated neutrophil influx
35 between the 2 groups following exposure to 200 ppb O3 for 2 hours with mild exercise
36 (Stenfors et al.. 2002). A more recent study found that atopic asthmatic subjects exhibited
37 an enhanced inflammatory response to O3 (400 ppb, 4 hours, with exercise) (Hernandez
38 etal.. 2010). This response was characterized by greater numbers of neutrophils, higher
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1 levels of IL-6, IL-8 and IL-1(3 and greater macrophage cell-surface expression of TLR4
2 and IgE receptors in induced sputum compared with healthy subjects. This study also
3 reported a greater increase in hyaluronan in atopic subjects and atopic asthmatics
4 compared with healthy subjects following O3 exposure. Animal studies have previously
5 reported that hyaluronic acid activates TLR4 signaling and results in AHR (see Section
6 5.3.5). Furthermore, levels of IL-10, a potent anti-inflammatory cytokine, were greatly
7 reduced in atopic asthmatics compared to healthy subjects. These results provide
8 evidence that innate immune and adaptive responses are different in asthmatics and
9 healthy subjects exposed to O3.
10 Eosinophils may be an important modulator of responses to O3 in asthma and allergic
11 airways disease. Eosinophils and associated proteins are thought to affect muscarinic
12 cholinergic receptors which are involved in vagally-mediated bronchoconstriction
13 (Mudway and Kelly. 2000). Studies described in Section 5.3.5 which demonstrated a key
14 role of eosinophils in O3-mediated AHR may be relevant to human allergic airways
15 disease which is characterized by airways eosinophilia (Yost et al.. 2005). Furthermore,
16 O3 exposure sometimes results in airways eosinophilia in allergic subjects or animal
17 models. For example, eosinophilia of the nasal and other airways was observed in
18 individuals with pre-existing allergic disease following O3 inhalation (270 and 400 ppb
19 O3, 2 hours, with exercise) (Vagaggini et al.. 2002; Peden etal.. 1995). Further, O3
20 exposure (0.5 ppm, 8 hours/day for 1-3 days) increased allergic responses, such as
21 eosinophilia and augmented intraepithelial mucosubstances, in the nasal airways of
22 ovalbumin (OVA)-sensitized rats (Wagner et al.. 2002). In contrast, Stenfors (2002) found
23 no stimulation of eosinophil influx measured in bronchial washes and BALF of mild
24 asthmatics following exposure to a lower concentration (200 ppb, 2 hours, with exercise)
25 ofO3.
26 The role of mast cells in O3-mediated asthma exacerbations has been investigated. Mast
27 cells are thought to play a key role in O3-induced airways inflammation, since airways
28 neutrophilia was decreased in mast cell-deficient mice exposed to O3 (Kleeberger et al..
29 1993). However, another study found that mast cells were not involved in the
30 development of increased bronchial reactivity in O3-exposed mice (Noviski et al.. 1999).
31 Nonetheless, mast cells release a wide variety of important inflammatory mediators
32 which may lead to asthma exacerbations (Stenfors et al.. 2002). A large increase in mast
33 cell number in bronchial submucosa was observed in non-asthmatics and a significant
34 decrease in mast cell number in bronchial epithelium was observed in mild asthmatics 6
35 hours following exposure to 200 ppb O3 for 2 hours during mild exercise (Stenfors et al..
36 2002). While these results point to an O3-mediated flux in bronchial mast cell
37 populations which differed between the non-asthmatics and mild asthmatics,
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1 interpretation of these findings is difficult. Furthermore, mast cell number did not change
2 in airway lavages in either group in response to O3 (Stenfors et al.. 2002)
3 Cytokine profiles in the airways have been investigated as an indicator of O3 sensitivity.
4 Differences in epithelial cytokine expression were observed in bronchial biopsy samples
5 in non-asthmatic and asthmatic subjects both at baseline and 6-hours postexposure to 200
6 ppb O3 for 2 hours (Bosson et al., 2003). The asthmatic subjects had a higher baseline
7 expression of IL-4 and IL-5 compared to non-asthmatics. In addition, expression of IL-5,
8 IL-8, GM-CSF, and ENA-78 in asthmatics was increased significantly following O3
9 exposure compared to non-asthmatics (Bosson et al.. 2003). Some of these (IL-4, IL-5
10 and GM-CSF) are Th2-related cytokines or neutrophil chemoattractants, and play a role
11 in IgE production, airways eosinophilia and suppression of Thl-cytokine production
12 (Bosson et al., 2003). These findings suggest a link between adaptive immunity and
13 enhanced responses of asthmatics to O3.
14 A further consideration is the compromised status of ELF antioxidants in disease states
15 such as asthma (Mudway and Kelly. 2000). This could possibly be due to ongoing
16 inflammation which causes antioxidant depletion or to abnormal antioxidant transport or
17 synthesis (Mudway and Kelly. 2000). For example, basal levels of AH2 were
18 significantly lower and basal levels of oxidized GSH and UA were significantly higher in
19 bronchial wash fluid and BALF of mild asthmatics compared with healthy control
20 subjects (Mudway et al.. 2001). Differences in ELF antioxidant content have also been
21 noted between species. These observations have led to the suggestion that the amount and
22 composition of ELF antioxidants, the capacity to replenish antioxidants in the ELF or the
23 balance between beneficial and injurious interactions between antioxidants and O3 may
24 contribute to O3 sensitivity, which varies between individuals and species (Mudway et
25 al.. 2006: Mudwav and Kelly. 2000: Mudwav et al.. 1999a). The complexity of these
26 interactions was demonstrated by a study in which a 2-hour exposure to 200 ppb O3,
27 while exercising, resulted in similar increases in airway neutrophils and decreases in
28 pulmonary function in both mild asthmatics and healthy controls, despite differences in
29 ELF antioxidant concentrations prior to O3 exposure (Mudway et al.. 2001). Further, the
30 O3-induced increase in oxidized GSH and decrease in AH2 observed in ELF of healthy
31 controls was not observed in mild asthmatics (Mudway et al.. 2001). While the authors
32 concluded that basal AH2 and oxidized GSH concentrations were not predictive of
33 responsiveness to O3, they also suggested that the increased basal UA concentrations in
34 the mild asthmatics may have played a protective role (Mudwav et al., 2001). Thus
3 5 compensatory mechanisms resulting in enhanced total antioxidant capacity may play a
36 role in modulating responses to O3.
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1 Collectively these older and more recent studies provide insight into mechanisms which
2 may contribute to enhanced responses of asthmatic and atopic individuals following O3
3 exposure. Greater airways inflammation and/or greater bronchial reactivity have been
4 demonstrated in asthmatics compared to non-asthmatics. This pre-existing inflammation
5 and altered baseline bronchial reactivity may contribute to the enhanced
6 bronchoconstriction seen in asthmatics exposed to O3. Furthermore, O3-induced
7 inflammation may contribute to O3-mediated AHR. An enhanced neutrophilic response
8 has been demonstrated in some asthmatics. A recent study in humans provided evidence
9 for differences in innate immune responses related to TLR4 signaling between asthmatics
10 and healthy subjects. Animal studies have demonstrated a role for eosinophil-derived
11 proteins in mediating the effects of O3. Since airways eosinophilia occurs in both allergic
12 humans and allergic animal models, this pathway may underlie the exacerbation of
13 allergic asthma by O3. In addition, differences have been noted in epithelial cytokine
14 expression in bronchial biopsy samples of healthy and asthmatic subjects. A Th2
15 phenotype, indicative of adaptive immune system activation and enhanced allergic
16 responses, was observed before O3 exposure and was increased by O3 exposure in
17 asthmatics. These findings support links between innate and adaptive immunity and
18 sensitivity to O3-mediated effects in asthmatics and allergic airways disease.
19 In addition to asthma and allergic diseases, obesity may alter responses to O3. While O3
20 is a trigger for asthma, obesity is a known risk factor for asthma (Shore. 2007). The
21 relationship between obesity and asthma is not well understood but recent investigations
22 have focused on alterations in endocrine function of adipose tissue in obesity. It is
23 thought that the increases in serum levels of factors produced by adipocytes (i.e.
24 adipokines) such as cytokines, chemokines, soluble cytokine receptors and energy
25 regulating hormones, may contribute to the relationship between obesity and asthma.
26 Some of these same mechanisms may be relevant to insulin resistant states such as
27 metabolic syndrome.
28 In a reanalysis of the data of Hazucha (2003). increasing body mass index in young
29 women was associated with increased O3 responsiveness, as measured by spirometry
30 following a 2-hour exposure to 500 ppb O3 while exercising (Bennett et al., 2007). In
31 several mouse models of obesity, airways were found to be innately more
32 hyperresponsive and responded more vigorously to acute O3 exposure than lean controls
33 (Shore. 2007). Pulmonary inflammatory and injury in response to O3 were also enhanced
34 (Shore. 2007). It was postulated that oxidative stress resulting from obesity-related
35 hyperglycemia could account for these effects (Shore. 2007). However, responses to O3
36 in the different mouse models are somewhat variable and depend on whether exposures
37 are acute or subacute. For example, diet-induced obesity augmented inflammation and
3 8 injury, as measured by BALF markers, and enhanced AHR in mice exposed acutely to O3
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1 (2 ppm, 3 hours) (Johnston et al.. 2008). In contrast, the inflammatory response following
2 sub-acute exposure to O3 was dampened by obesity in a different mouse model (0.3 ppm,
3 72 hours) (Shore et al.. 2009).
5.4.2.3 Nutritional Status
4 Many investigations have focused on antioxidant deficiency and supplementation as
5 modulators of O3-mediated effects. One study in mice found that vitamin A deficiency
6 enhanced lung injury induced by exposure to 0.3 ppm O3 for 72 hours (Paquette et al..
7 1996). Ascorbate deficiency was shown to increase the effects of acute (0.5-1 ppm for 4
8 hours), but not subacute (0.2-0.8 ppm for 7 days), O3 exposure in guinea pigs (Kodavanti
9 et al.. 1995; Slade et al.. 1989). Supplementation with AH2 and a-TOH was protective in
10 healthy adults who were on an AH2-deficient diet and exposed to 400 ppb O3 for 2 hours
11 while exercising (Samet etal.. 2001). In this study, the protective effect consisted of a
12 smaller reduction in FEVi following O3 exposure (Samet et al.. 2001). However the
13 inflammatory response (influx of neutrophils and levels of IL-6) measured in BALF 1
14 hour after O3 exposure was not different between supplemented and non-supplemented
15 subj ects (Samet etal.. 2001). Other investigators found that AH2 and a-TOH
16 supplementation failed to ameliorate the pulmonary function decrements or airways
17 neutrophilia observed in humans exposed to 200 ppb O3 for 2 hours (Mudway et al..
18 2006). It was suggested that supplementation may be ineffective in the absence of
19 antioxidant deficiency (Mudway et al.. 2006).
20 In asthmatic adults, these same dietary antioxidants reduced O3-induced bronchial
21 hyperresponsiveness (120 ppb, 45 min, with exercise) (Trenga et al.. 2001). Furthermore,
22 supplementation with AH2 and a-tocopherol protected against pulmonary function
23 decrements and nasal inflammatory responses which were associated with high levels of
24 ambient O3 in asthmatic children living in Mexico City (Sienra-Monge et al.. 2004;
25 Romieu et al.. 2002). Similarly, supplementation with ascorbate, a-tocopherol and
26 (3-carotene improved pulmonary function in Mexico City street workers (Romieu et al..
27 1998a).
28 Protective effects of supplementation with a-tocopherol alone have not been observed in
29 humans experimentally exposed to O3 (Mudway and Kelly. 2000). Alpha-TOH
30 supplementation also failed to protect against O3-induced effects in animal models of
31 allergic rhinosinusitis and lower airways allergic inflammation (rats, 1 ppm O3 for 2
32 days) (Wagner et al.. 2007). However, protection in these same animal models was
33 reported using y-TOH supplementation (Wagner et al.. 2009; Wagner et al.. 2007). Other
34 investigators found that a-TOH deficiency led to an increase in liver lipid peroxidation
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1 (rats, 0.3 ppm 3 hours/day for 7 months) (Sato et al., 1980) and a drop in liver a-TOH
2 levels following O3 exposure (mice, 0.5 ppm, 6 hours/day for 3 days) (Vasu etal.. 2010).
3 A recent study used a-TOH transfer protein null mice as a model of a-TOH deficiency
4 and demonstrated an altered adaptive response of the lung genome to O3 exposure (Vasu
5 etal., 2010). Taken together, these studies provide evidence that the tocopherol system
6 modulates Os-induced responses.
5.4.2.4 Lifestage
7 Responses to O3 are modulated by factors associated with lifestage. On one end of the
8 lifestage spectrum is aging. The spirometric response to O3 appears to be lost in humans
9 as they age, as was demonstrated in two studies involving exposures of exercising human
10 subjects to 420-450 ppb O3 for 1.5-2 hours (Hazucha et al., 2003; Drechsler-Parks.
11 1995). In mice, physiological responses to O3 (600 ppb, 2 hours) were diminished with
12 age (Hamade et al., 2010). Mechanisms accounting for this effect have not been well-
13 studied but could include altered number and sensitivity of receptors or altered signaling
14 pathways involved in neural reflexes.
15 On the other side of the lifestage spectrum is pre/postnatal development. Critical
16 windows of development during the pre/postnatal period are associated with an enhanced
17 sensitivity to environmental toxicants. Adverse birth outcomes and developmental
18 disorders may occur as a result.
19 Adverse birth outcomes may result from stressors which impact transplacental oxygen
20 and nutrient transport by a variety of mechanisms including oxidative stress, placental
21 inflammation and placental vascular dysfunction (Kannan et al. 2006). These
22 mechanisms may be linked since oxidative/ nitrosative stress is reported to cause vascular
23 dysfunction in the placenta (Myatt et al.. 2000). As described in Section 7.4, systemic
24 inflammation and oxidative/nitrosative stress and modification of innate and adaptive
25 immunity are key events underlying the health effects of O3 and as such they may
26 contribute to adverse birth outcomes. An animal toxicology study showing that exposure
27 to 2 ppm O3 led to anorexia (Kavlock et al., 1979) (see Section 7.4.2) in exposed rat
28 dams provide an additional mechanism by which O3 exposure could lead to diminished
29 transplacental nutrient transport. Disturbances of the pituitary-adrenocortico-placental
30 system (Ritz et al.. 2000) may also impact normal intrauterine growth and development.
31 Further, restricted fetal growth may result from pro-inflammatory cytokines which limit
32 trophoblast invasion during the early stages of pregnancy (Hansen et al.. 2008). Direct
33 effects on maternal health, such as susceptibility to infection, and on fetal health, such as
34 DNA damage, have also been proposed as mechanisms underlying adverse birth
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1 outcomes (Ritz et al., 2000). In addition to restricted fetal growth, preterm birth may
2 contribute to adverse birth outcomes. Preterm birth may result from the development of
3 premature contractions and/or premature rupture of membranes as well as from disrupted
4 implantation and placentation which results in suboptimal placental function (Darrow et
5 al., 2009; Ritz et al., 2000). Genetic mutations are thought to be an important cause of
6 placental abnormalities in the first trimester, while vascular alterations may be the main
7 cause of placental abnormalities in later trimesters (Jalaludin et al.. 2007). Ozone-
8 mediated systemic inflammation and oxidative stress/nitrosative stress may possibly be
9 related to these effects although there is no firm evidence.
10 Enhanced sensitivity to environmental toxicants during critical windows of development
11 may also result in developmental disorders. For example, normal migration and
12 differentiation of neural crest cells are important for heart development and are
13 particularly sensitive to toxic insults (Ritz et al.. 2002). Further, immune dysregulation
14 and related pathologies are known to be associated with pre/postnatal environmental
15 exposures (Dietert et al.. 2010). Ozone exposure is associated with developmental effects
16 in several organ systems. These include neurobehavioral changes which could reflect
17 O3 's effects on CNS plasticity or the hypothalamic-pituitary axis (Auten and Foster. In
18 Press) (see Section 7.4.9).
19 The majority of developmental effects due to O3 have been described for the respiratory
20 system (see Section 7.2.3 and 7.4.8). Since its growth and development take place during
21 both the prenatal and early postnatal periods, both prenatal and postnatal exposures to O3
22 have been studied. Maternal exposure to 0.4-1.2 ppm O3 during gestation resulted in
23 developmental health effects in the RT of mice (Sharkhuu et al.. 2011). Recent studies
24 involving postnatal exposure to O3 have focused on differences between developing and
25 adult animals in antioxidant defenses, respiratory physiology and sensitivity to cellular
26 injury (Auten and Foster. In Press). In particular, one set of studies in infant rhesus
27 monkeys exposed to 0.5 ppm O3 intermittently over 5 months has identified numerous
28 O3-mediated perturbations in the developing lung and immune system (Plopperetal..
29 2007). These investigations were prompted by the dramatic rise in the incidence of
30 childhood asthma and focused on the possible role of O3 and allergens in promoting
31 remodeling of the epithelial-mesenchymal trophic unit during postnatal development of
32 the tracheobronchial airway wall. These and other studies have focused on mechanisms,
33 such as lung structural changes, antigen sensitization, interaction with nitric oxide
34 signaling, altered airway afferent innervation and loss of alveolar repair capacity, by
35 which early O3 exposure could lead to asthma pathogenesis or exacerbations in later life
36 (Auten and Foster. In Press). Further, a recent study demonstrated that maternal exposure
37 to particulate matter (PM) resulted in augmented lung inflammation, airway epithelial
3 8 mucous metaplasia and AHR in young mice exposed chronically and intermittently to 1
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1 ppm O3 (Auten et al.. 2009). Early life exposure to O3 has also been found to modulate
2 pulmonary and systemic innate immunity later in life in the infant rhesus monkey model
3 (Maniar-Hew et al.. 2011).
5.4.2.5 Attenuation of Responses
4 In responsive individuals, a striking degree of response attenuation occurred following
5 repeated daily exposures to O3. Generally, the young O3 responder was no longer
6 responsive on the fourth or fifth day of consecutive daily O3 exposure (200-500 ppb O3
7 for 2-4 hours) and required days to weeks of non-exposure in order for the subject to
8 regain O3 responsiveness (Christian et al.. 1998; Devlin et al.. 1997; Linn et al.. 1982a;
9 Horvath et al.. 1981; Hackney et al.. 1977). This phenomena has been reported for both
10 lung function and symptoms such as upper airway irritation, nonproductive cough,
11 substernal discomfort and pain upon deep inspiration (Linnet al.. 1982a; Horvath et al..
12 1981; Hackney et al.. 1977). Repeated daily exposures also led to an attenuation of the
13 sRaw response in exercising human subjects exposed for 4 hours to 200 ppb O3
14 (Christian et al.. 1998) and to a dampened AHR response compared with a single day
15 exposure in exercising human subjects exposed for 2 hours to 400 ppb O3 (Dimeo et al..
16 1981). However, one group reported persistent small airway dysfunction despite
17 attenuation of the FEVi response on the third day of consecutive O3 exposure (250 ppb,
18 2 hours, with exercise) (Frank et al.. 2001).
19 Studies in rodents also indicated an attenuation of the physiologic response measured by
20 breathing patterns and tidal volume following five consecutive days of exposure to 0.35-1
21 ppm O3 for 2.25 hours (Tepper et al.. 1989). Attenuation of O3-induced bradycardic
22 responses, which also result from activation of neural reflexes, has been reported in
23 rodents (0.5-0.6 ppm O3, 2-6 h/dy, 3-5 days (Hamade and Tankersley. 2009; Watkinson et
24 al.. 2001).
25 Multi-day exposure to O3 has been found to decrease some markers of inflammation
26 compared with a single day exposure (Christian et al.. 1998; Devlin et al.. 1997). For
27 example, in human subjects exposed for 4 hours to 200 ppb O3 during moderate exercise,
28 decreased numbers of BAL neutrophils and decreased levels of BALF fibronectin and IL-
29 6 were observed after 4 days of consecutive exposure compared with responses after 1
30 day (Christian et al.. 1998). Results indicated an attenuation of the inflammatory response
31 in both proximal airways and distal lung. However markers of injury, such as lactate
32 dehydrogenase (LDH) and protein in the BALF, were not attenuated in this study
33 (Christian et al.. 1998). Other investigators found that repeated O3 exposure (200 ppb O3
34 for 4 hours on 4 consecutive days with intermittent exercise) resulted in increased
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1 numbers of neutrophils in bronchial mucosal biopsies despite decreased BAL
2 neutrophilia (Torres et al.. 2000). Other markers of inflammation, including BALF protein
3 and IL-6 remained elevated following the multi-day exposure (Torres et al.. 2000).
4 In rats, the increases in BALF levels of proteins, fibronectin, IL-6 and inflammatory cells
5 observed after one day of exposure to 0.4 ppm O3 for 12 hours were no longer observed
6 after 5 consecutive days of exposure (Van Bree et al., 2002). A separate study in rats
7 exposed to 0.35-1 ppm O3 for 2.25 hours for 5 consecutive days demonstrated a lack of
8 attenuation of the increase in BALF protein, persistence of macrophages in the
9 centriacinar region and histological evidence of progressive tissue injury (Tepper et al..
10 1989). Findings that injury, measured by BALF markers or by histopathology, persist in
11 the absence of BAL neutrophila or pulmonary function decrements suggested that
12 repeated exposure to O3 may have serious long-term consequences such as airway
13 remodeling. In particular, the small airways were identified as a site where cumulative
14 injury may occur (Frank et al., 2001).
15 Some studies examined the recovery of responses which were attenuated by repeated O3
16 exposure. In a study of humans undergoing heavy intermittent exercise who were
17 exposed for 2 hours to 400 ppb O3 for five consecutive days (Devlin et al.. 1997).
18 recovery of the inflammatory responses which were diminished by repeated exposure
19 required 10-20 days following the exposure (Devlin et al.. 1997). In an animal study
20 conducted in parallel (Van Bree et al.. 2002). full susceptibility to O3 challenge following
21 exposure to O3 for five consecutive days required 15-20 days recovery.
22 Several mechanisms have been postulated to explain the attenuation of responses
23 observed in human subjects and animal models following repeated exposure to O3. First,
24 the upregulation of antioxidant defenses (or conversely, a decrease in critical O3-reactive
25 substrates) may protect against O3-mediated adverse effects. Increases in antioxidant
26 content of the BALF have been demonstrated in rats exposed to 0.25 and 0.5 ppm O3 for
27 several hours on consecutive days (Devlin et al.. 1997; Wiester et al.. 1996a; Tepper et
28 al.. 1989). Second, IL-6 was demonstrated to be an important mediator of attenuation in
29 rats exposed to 0.5 ppm for 4 hours on two consecutive days (McKinney et al.. 1998).
30 Third, a protective role for increases in mucus producing cells and mucus concentrations
31 in the airways has been proposed (Devlin et al.. 1997). Fourth, epithelial hyperplasia or
32 metaplasia may decrease susceptibility to subsequent O3 challenge (Carey et al.. 2007;
33 Harkema et al.. 1987a; Harkema et al.. 1987b). These morphologic changes have been
34 observed in nasal and lower airways in monkeys exposed chronically to 0.15-0.5 ppm O3.
35 Although there is some evidence to support these possibilities, there is no consensus on
36 mechanisms underlying response attenuation. Recent studies demonstrating that O3
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1 activates TRP receptors suggest that modulation of TRP receptor number or sensitivity by
2 repeated O3 exposures may also contribute to the attenuation of responses.
5.4.2.6 Co-Exposures with Particulate Matter
3 Numerous studies have investigated the effects of co-exposure to O3 and PM because of
4 the prevalence of these pollutants in ambient air. Results are highly variable and depend
5 on whether exposures are simultaneous or sequential, the type of PM employed and the
6 endpoint examined. Additive and interactive effects have been demonstrated. For
7 example, simultaneous exposure to O3 (120 ppb for 2 hours at rest) and concentrated
8 ambient particles (CAPs) in human subjects resulted in a diminished systemic IL-6
9 response compared with exposure to CAPs alone (Urch et al.. 2010). However, exposure
10 to O3 alone did not alter blood IL-6 levels (Urchet al., 2010). The authors provided
11 evidence that O3 mediated a switch to shallow breathing which may have accounted for
12 the observed antagonism (Urchet al.. 2010). Further, simultaneous exposure to O3 (114
13 ppb for 2 hours at rest) and CAPs but not exposure to either alone, resulted in increased
14 diastolic blood pressure in human subjects (Fakhri et al., 2009). Mechanisms underlying
15 this potentiation of response were not explored. In some strains of mice, pre-exposure to
16 O3 (0.5 ppm for 2 hours) modulated the effects of carbon black PM on heart rate, HRV
17 and breathing patterns (Hamade and Tankersley. 2009). Another recent study in mice
18 demonstrated that treatment with carbon nanotubes followed 12 hours later by O3
19 exposure (0.5 ppm for 3 hours) resulted in a dampening of some of the pulmonary effects
20 of carbon nanotubes measured as markers of inflammation and injury in the BALF (Han
21 et al.. 2008). Lastly, Harkema et al. (2005) found that epithelial and inflammatory
22 responses in the airways of rats were enhanced by co-exposure to O3 (0.5 ppm for 3 days)
23 and LPS (used as a model of biogenic PM) or to O3 (1 ppm for 2 days) and OVA (used as
24 a model of an aeroallergen). Many of the demonstrated responses were more-than-
25 additive. Overall, these findings are hard to interpret but demonstrate the complexity of
26 responses following combined exposure to PM and O3.
5.4.2.7 Summary
27 Collectively, these older and more recent studies provide evidence for mechanisms which
28 may underlie the variability in responsiveness seen among individuals (Figure 5-10).
29 Certain functional genetic polymorphisms, pre-existing conditions and diseases,
30 nutritional status, lifestage and co-exposures contribute to altered risk of O3-induced
31 effects. Attenuation of responses may also be important, but it is incompletely
32 understood, both in terms of the pathways involved and the resulting consequences.
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Dosimetric factors
Nutritional status
Life stage
Attenuation factors
Co-exposures
Ozone + Respiratory Tract
Gene-environment interactions
Pre-existing diseases/conditions
COPD/smoking status
Asthma/allergic airways disease
Obesity/metabolic syndrome
\
Formation of secondary oxidation products
Activation
of neural
reflexes
Initiation of
inflammation
Y7
Systemic inflammation and
oxidative/nitrosative stress
Sensitization
of bronchial
smooth muscle
Respiratory System Effects
Extrapulmonary Effects
Obesity/
Metabolic Stress
Lifestage
Attenuation
factors
Figure 5-10 Factors which contribute to the interindividual variability in
responses resulting from inhalation exposure to ozone.
5.5 Species Homology and Interspecies Sensitivity
1 The previous O3 AQCDs discussed the suitability of animal models for comparison with
2 human O3 exposure and concluded that the acute and chronic functional responses of
3 laboratory animals to O3 appear qualitatively homologous to human responses. Thus,
4 animal studies can provide important data in determining cause-effect relationships
5 between exposure and health outcome that would be impossible to collect in human
6 studies. Still, care must be taken when comparing quantitative dose-response
7 relationships in animal models to humans due to obvious interspecies differences. This
8 section will describe basic concepts in species homology concerning both dose and
9 response to O3 exposure. This will not be a quantitative extrapolation of doses where O3
10 effects have been observed. Overall, there have been few new publications examining
11 interspecies differences in dosimetry and response to O3 since the last AQCD. These
12 studies do not overtly change the conclusions discussed in the previous document.
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5.5.1 Dosimetry
1 As discussed in Section 5.2.1, O3 uptake depends on complex interactions between RT
2 morphology, breathing route, rate, and depth, physicochemical properties of the gas,
3 physical processes of gas transport, as well as the physical and chemical properties of the
4 ELF and tissue layers. Understanding differences in these variables between humans and
5 experimental animals is important to interpreting delivered doses in animal and human
6 toxicology studies.
7 Physiological and anatomical differences exist between experimental species. The
8 structure of the URT is vastly different between rodents and humans and scales according
9 to body mass. The difference in the cross-sectional shape and size of the nasal passages
10 affects bulk airflow patterns such that major airflow streams are created. The nasal
11 epithelium is lined by squamous, respiratory, or olfactory cells, depending on location.
12 The differences in airflow patterns in the URT mean that not all nasal surfaces and cell
13 types receive the same exposure to inhaled O3 leading to differences in local absorption
14 and potential for site-specific tissue damage. The morphology of the LRT also varies
15 within and among species. Rats and mice do not possess respiratory bronchioles;
16 however, these structures are present in humans, dogs, ferrets, cats, and monkeys.
17 Respiratory bronchioles are abbreviated in hamsters, guinea pigs, sheep, and pigs. The
18 branching structure of the ciliated bronchi and bronchioles also differs between species
19 from being a rather symmetric and dichotomous branching network of airways in humans
20 and primates to a more monopodial branching network in other mammals. In addition,
21 rodents have fewer terminal bronchioles due to a smaller lung size compared to humans
22 or canines (TvIcBride. 1992). The cellular composition in the pulmonary region is similar
23 across mammalian species; at least 95% of the alveolar epithelial tissue is composed of
24 Type I cells. However, significant differences exist between species in the number and
25 type of cells in the TB airways. Differences also exist in breathing route and rate.
26 Primates are oronasal breathers, while rodents are obligate nasal breathers. Past studies of
27 the effect of body size on resting oxygen consumption also suggest that rodents inhale
28 more volume of air per lung mass than primates. These distinctions as well as differences
29 in nasal structure between primates and rodents could affect the amount of O3 uptake.
30 As O3 absorption and activity relies on ELF antioxidant substances as described in
31 Section 5.2.3, variability in antioxidant concentrations and metabolism between species
32 may affect dose and O3-induced health outcomes. The thickness of the ELF in the TB
33 airways varies among species. Mercer et al. (1992) found that the human ELF thickness
34 in bronchi and bronchioles was 6.9 and 1.8 (im, respectively, compared to 2.6 and 1.9 (im
3 5 for the same locations in the rat. Guinea pigs and mice have a lower basal activity of
36 GSH transferase and GSH peroxidase, and lower a-TOH levels in the lung compared to
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1 rats (Ichinose et al., 1988; Sagai et al., 1987). Nasal lavage fluid analysis shows that
2 humans have a higher proportion of their nasal antioxidants as UA and low levels of AH2
3 whereas mice, rats, or guinea pigs have high levels of AH2 and undetectable levels of UA
4 (Figure 5-1 la). GSH is not detected in the nasal lavage of most of these species, but is
5 present in monkey nasal lavage. Guinea pigs and rats have a higher antioxidant to protein
6 ratio in nasal lavage and BALF than humans (Hatch. 1992). The BALF profile differs
7 from the nasal lavage (Figure 5-1 Ib). Humans have a higher proportion of GSH and less
8 AH2 making up their BALF content compared to the guinea pigs and rats (Slade et al..
9 1993; Hatch. 1992). Similar to the nose, rats have the highest antioxidant to protein mass
10 ratio found in BALF (Slade etal.. 1993). Antioxidant defenses also vary with age
11 (Servais et al.. 2005) and exposure history (Duanetal.. 1996). Duan et al. (1996; 1993)
12 reported that differences in antioxidant levels between species and lung regions did not
13 appear to be the primary factor in O3 induced tissue injury. However, a close association
14 between site-specific O3 dose, the degree of epithelial injury, and reduced glutathione
15 depletion was later revealed in monkeys (Plopperetal.. 1998).
16 Humans and animals are similar in the pattern of regional O3 dose distribution. As
17 discussed for humans in Section 5.2.2, O3 flux to the air-liquid interface of the ELF
18 slowly decreases distally in the TB region and then rapidly decreases distally in the
19 alveolar region (Miller et al.. 1985). Modeled tissue dose in the human RT, representing
20 O3 flux to the liquid-tissue interface, is very low in the trachea, increases to a maximum
21 in the CAR, and then rapidly decreases distally in the alveolar region (Figure 5-12).
22 Similar patterns of O3 tissue dose profiles normalized to inhaled O3 concentration were
23 predicted for rat, guinea pig, and rabbit (Miller etal.. 1988; Overton etal.. 1987) (Figure
24 5-12a). Overton et al. (1987) modeled rat and guinea pig O3 dose distribution and found
25 that after comparing two different morphometrically based anatomical models for each
26 species, considerable difference in predicted percent RT and alveolar region uptakes were
27 observed. This was due to the variability between the two anatomical models in airway
28 path distance to the first alveolated duct. As a result, the overall dose profile was similar
29 between species however the O3 uptake efficiency varied due to RT size and path length
30 (Section 5.2.2). A similar pattern of O3 dose distribution was measured in monkeys
31 exposed to 0.4 and 1.0 ppm 18O3 (Plopper et al.. 1998) (Figure 5-12b). Less 18O was
32 measured in the trachea, proximal bronchus, and distal bronchus than was observed in the
33 respiratory bronchioles. Again indicating the highest concentration of O3 tissue dose to
34 be localized to the CAR, which are the respiratory bronchioles in nonhuman primates. In
35 addition, the lowest 18O detected in the RT was in the parenchyma (i.e. alveolar region),
36 mimicking the rapid decrease in tissue O3 dose predicted by models for the alveolar
37 regions of humans and other animals.
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b.
Rat
Guinea pig
Human
1
1
a
D Ascorbic acid
• Uric acid
D Glutathione
Rat
Guinea pig
Human
200 400 600 800
Antioxidant/ Protein, nanomoles / gram
1000
n
0 50 100 150 200 250
Antioxidant/ Protein, nanomoles / gram
Source: Adapted with permission from CRC Press, Inc. (Sladeet al.. 1993: Hatch. 1992)
Figure 5-11 Species comparison of antioxidant / protein ratios of: (a) nasal
lavage fluid and, (b) bronchoalveolar lavage fluid.
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a. ^
E 1C)
"~«
o
n
E
(0
c
E
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o
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0)
=1
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to
10'
10
b.
05 50
1
40
30
20
•51
OS
c
c
o
o
£ 10
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0)
O)
I •
Human
Rat ----
Guinea Pig
f (bpm)
15.0
66.0
60.9
Rabbit
_.^.. 13.20 38.8
(No absorption in the URT)
TB
I— Zone 0
Order 0
Generation 0
Generation 0
6
10
10
5
7
12
12
14
14
6 7
91011
151617
161718
**
12
19
20
13
21
22
8 Rabbit
14 Guinea Pig
23 Rat
23 Human
0.4ppmO3
1.0ppmO3
TRACHEA PROXIMAL DISTAL RESPIRATORY PARENCHYMA
BRONCHUS BRONCHUS BRONCHIOLE
Source: Panel (a) U.S. EPAQ996a) (b) Plopper et al. (1998)
Figure 5-12 Humans and animals are similar in the regional pattern of Os tissue
dose distribution. Panel (a) presents the predicted tissue dose of
Os (as ug of Os per cm2 of segment surface area per min,
standardized to a trachea! Os value of 1 ug/m3) for various regions
of the rabbit, guinea pig, rat, and human RT. TB = tracheobronchial
region, A = alveolar region. Panel (b) presents a comparison of
excess 18O in the five regions of the TB airways of rhesus monkeys
exposed to Os for 2h. *p<0.05 comparing the same Os
concentration across regions. **p<0.05 comparing different Os
concentrations in the same region.
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1
2
3
4
5
6
7
8
9
10
11
12
Humans and animal models are similar in the pattern of regional O3 dose, but absolute
values differ. Hatch et al. (1994) reported that exercising humans exposed to oxygen-18
labeled O3 (400 ppb) accumulated 4-5 times higher concentrations of O3 reaction product
in BAL cells, surfactant and protein fractions compared to resting rats similarly exposed
(0.4 ppm) (Figure 5-13). It was necessary to expose resting rats to 2 ppm O3 to achieve
the same BALF accumulation of 18O reaction product that was observed in humans
exposed to 400 ppb with intermittent heavy exercise (MV ~60 L/min). The concentration
of 18O reaction product in BALF paralleled the accumulation of BALF protein and
cellular effects of the O3 exposure observed such that these responses to 2.0 ppm O3 were
similar to those of the 400 ppb O3 in exercising humans. This suggests that animal data
obtained in resting conditions would underestimate the dose to the RT and presumably
the resultant risk of effect for humans.
60
50
40
9 30
8
UJ 20
10
Source: Hatch et al. (1994)
BAL Cells
BAL HSP
BAL HSS
Lavaged Lung
Exercising Human
(0.4 ppm, 2 hours)
Resting F-344 Rat
(0.4 ppm, 2 hours)
Resting F-344 Rat
(2.0 ppm, 2 hours)
Figure 5-13 Oxygen-18 incorporation into different fractions of BALF from
humans and rats exposed to 0.4 and 2.0 ppm 18O3.The excess 18O
in each fraction is expressed relative to the dry weight of that
fraction. Fractions assayed include cells, high speed pellet (HSP),
high speed supernatant (HSS), and lavaged lung homogenates.
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1 Recently, a quantitative comparison of O3 transport in the airways of rats, dogs, and
2 humans was conducted using a three-compartment airways model, based on upper and
3 lower airway casts and mathematical calculation for alveolar parameters (Tsujino etal..
4 2005). This model examined how interspecies anatomical and physiological differences
5 affect intra-airway O3 concentrations and the amount of gas absorbed. The model was
6 designed as cylindrical tubes with constant volume and one-dimensional gas movement
7 and no airway branching patterns. Peak, real-time, and mean O3 concentrations were
8 higher in the upper and lower airways of humans compared to rats and dogs, but lowest
9 in the alveoli of humans. The amount of O3 absorbed was lowest in humans when
10 normalized by body weight. The intra-airway concentration decreased distally in all
11 species. Sensitivity analysis demonstrated that VT, fB, and upper and lower airways
12 surface area had a significant impact on model results. The model is limited in that it did
13 not account for chemical reactions in the ELF or consider gas diffusion as a driving force
14 for O3 transport. Also, the model was run at a respiratory rate of 16/min simulating a
15 resting individual, however exercise may cause a further deviation from animal models as
16 was seen in Hatch et al. (1994).
17 Overall, animal models exhibit qualitatively similar patterns of O3 net and tissue dose
18 distribution with the largest tissue dose delivered to the CAR. However, due to
19 anatomical and biochemical RT differences the absolute values of O3 dose delivered
20 differs. Past results suggest that animal data obtained in resting conditions would
21 underestimate the dose to the RT and presumably the resultant risk of effect for humans,
22 especially for humans during exercise.
5.5.2 Homology of Response
23 Risk of heath effects from O3 varies between and within species, as well as between
24 endpoints. Rodents appear to have a slightly higher tachypneic response to O3 and are
25 less sensitive to changes in pulmonary function test than humans (U.S. EPA. 1996a).
26 However, rats experience attenuation of pulmonary function and tachypneic ventilatory
27 responses, similar to humans (Wiester et al.. 1996a). Hatch et al. (1986) reported that
28 guinea pigs were the most responsive to O3-induced inflammatory cell and protein influx.
29 Rabbits were the least responsive and rats, hamsters, and mice were intermediate
30 responders. Further analysis of this study by Miller et al. (1988) found that the protein
31 levels in guinea pigs increased more rapidly with predicted pulmonary tissue dose than in
32 rats and rabbits. Alveolar macrophages isolated from guinea pigs and humans mounted
33 similar qualitative and quantitative cytokine responses to in vitro O3 (0.1-1.0 ppm for 60
34 minutes) exposure (Arsalane et al.. 1995).
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1 Also, because of their higher body surface to volume ratio, rodents can rapidly lower
2 body temperature during exposure leading to lowered O3 dose and toxicity CWatkinson et
3 al.. 2003; Iwasaki et al., 1998; SladeetaL 1997). In addition to lowering the O3 dose to
4 the lungs, this hypothermic response may cause: (1) lower metabolic rate, (2) altered
5 enzyme kinetics, and (3) altered membrane function. The thermoregulatory mechanisms
6 also may affect disruption of heart rate which may lead to: (1) decreased cardiac output,
7 (2) lowered blood pressure, and (3) decreased tissue perfusion (Watkinson et al., 2003).
8 These responses have not been observed in humans except at very high exposures, thus
9 further complicating extrapolation of effects from animals to humans.
10 Recently, the three-dimensional detail of the nasal passages of immature Rhesus macaque
11 monkeys was analyzed for developing predictive dosimetry models and exposure-dose-
12 response relationships (Carey et al., 2007). In doing so the authors reported that the
13 relative amounts of the five epithelial cell types in the nasal airways of monkeys remains
14 consistent between infancy and adulthood (comparing to (Gross et al., 1987; Gross et al.,
15 1982). Ozone exposures (0.5 ppm, 8 h/day under acute [5 days] and episodic conditions
16 [5 replicates of the acute paradigm spaced a week apart]) confirmed that the ciliated
17 respiratory and transitional epithelium were the most sensitive cell types in the nasal
18 cavity to O3 exposure, showing 50-80% decreases in epithelial thickness and epithelial
19 cell volume. The character and location of nasal lesions resulting from O3 exposure were
20 similar between adult and infant monkeys similarly exposed. However, infant monkeys
21 did not undergo nasal airway epithelial remodeling or adaptation that occurs in adult
22 animals and they may develop persistent necrotizing rhinitis following episodic longer-
23 term exposures.
24 To further understand the genetic basis for age-dependent differential response to O3,
25 adult (15 week old) and neonatal (15-16 day old) mice from 8 genetically diverse strains
26 were examined for O3-induced (0.8 ppm for 5 hours) pulmonary injury and lung
27 inflammation (Vancza et al.. 2009). Ozone exposure increased polymorphonuclear
28 leukocytes (PMN) influx in all strains of neonatal mice tested, but significantly greater
29 PMNs occurred in neonatal compared to adult mice for only some sensitive strains,
30 suggesting a genetic background effect. This strain difference was not due to differences
31 in delivered dose of O3 to the lung, evidenced by 18O lung enrichment. The sensitivity of
32 strains for O3-induced increases in BALF protein and PMNs was different for different
33 strains of mice suggesting that genetic factors contributed to heightened responses.
34 Interestingly, adult mice accumulated more than twice the levels of 18O reaction product
35 of O3 than corresponding strain neonates. Thus, it appeared that the infant mice showed a
36 two- to threefold higher response than the adults when expressed relative to the
37 accumulated O3 reaction product in their lungs. The apparent decrease in delivered O3
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1 dose in neonates could be a result of a more rapid loss of body temperature in infant
2 rodents incident to maternal separation and chamber air flow.
3 Further, O3-induced injury and inflammation responses are variable between species. For
4 example, Dormans et al. (1999) found that rats, mice, and guinea pigs all exhibited
5 O3-induced (0.2 - 0.4 ppm for 3-56 days) inflammation; however, guinea pigs were the
6 most sensitive with respect to alveolar macrophage elicitation and pulmonary cell density
7 in the centriacinar region. Mice were the most sensitive to bronchiolar epithelial
8 hypertrophy and biochemical changes (e.g. LDH, glutathione reductase, glucose-6-
9 phosphate dehydrogenase activity), and had the slowest recovery from O3 exposure. All
10 species displayed increased collagen in the ductal septa and large lamellar bodies in Type
11 II pneumocytes at the longest exposure and highest concentration; whereas this response
12 occurred in the rat and guinea pig at lower O3 levels (0.2 ppm) as well. Overall, the
13 authors rated mice as most sensitive, followed by guinea pigs, then rats (Dormans et al..
14 1999). Rats were also less sensitive to epithelial necrosis and inflammatory responses
15 from O3 (1.0 ppm for 8 hours) than monkeys and ferrets, which manifested a similar
16 response (Sterner-Kock et al.. 2000). These data suggest that ferrets may be a good
17 animal model for O3-induced airway effects due to the similarities in pulmonary structure
18 between primates and ferrets. However, this study provided no dose metric and, it is
19 possible that some of these differences may be attributable to disparate total inhaled dose
20 or local organ dose.
5.5.3 Summary
21 In summary, for all species there are limitations that must be considered when attempting
22 to extrapolate to human O3 exposures. Rats required 4-5 times higher exposure to O3 to
23 achieve comparable increases in BALF protein and PMNs to exercising humans. New
24 studies have shown that varied O3 response in different mouse strains was not due to
25 differences in delivered dose of O3 to the lung but more likely genetic sensitivity, and that
26 infant mice show greater toxicity relative to their smaller lung dose than adults. Even
27 though interspecies differences limit quantitative comparison between species, the acute
28 and chronic functional responses of laboratory animals to O3 appear qualitatively
29 homologous to those of the human making them a useful tool in determining mechanistic
30 and cause-effect relationships with O3 exposure.
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5.6 Chapter Summary
1 Ozone is a highly reactive gas and a powerful oxidant with a short half-life. Both O3
2 uptake and responses are dependent upon the formation of secondary reaction products in
3 the ELF; however more complex interactions occur. Uptake in humans at rest is 80-95%
4 efficient and it is influenced by a number of factors including RT morphology, breathing
5 route, frequency, and volume, physicochemical properties of the gas, physical processes
6 of gas transport, as well as the physical and chemical properties of the ELF and tissue
7 layers. The primary uptake site of O3 delivery to the lung epithelium is believed to be the
8 CAR, however changes in a number of factors (e.g. physical activity) can alter the
9 distribution of O3 uptake in the RT. Ozone uptake is chemical reaction-dependent and the
10 substances present in the ELF appear in most cases to limit interaction of O3 with
11 underlying tissues and to prevent penetration of O3 distally into the RT. Still, reactions of
12 O3 with soluble ELF components or plasma membranes result in distinct products, some
13 of which are highly reactive and can injure and/or transmit signals to RT cells.
14 Thus, in addition to contributing to the driving force for O3 uptake, formation of
15 secondary oxidation products initiates pathways that provide the mechanistic basis for
16 health effects which are described in detail in Chapters 6 and 7 and which involve the RT
17 as well as extrapulmonary systems. These pathways include activation of neural reflexes,
18 initiation of inflammation, alteration of epithelial barrier function, sensitization of
19 bronchial smooth muscle, modification of innate and adaptive immunity, airways
20 remodeling, and systemic inflammation and oxidative/nitrosative stress. With the
21 exception of airways remodeling, these pathways have been demonstrated in both
22 animals and human subjects in response to the inhalation of O3.
23 Both dosimetric and mechanistic factors contribute to the understanding of
24 interindividual variability in responses to O3. Interindividual variability is influenced by
25 variability in RT volume and thus surface area, certain genetic polymorphisms, pre-
26 existing conditions and disease, nutritional status, lifestages, attenuation, and co-
27 exposures. Some of these factors are also influential in understanding species homology
28 and sensitivity. Qualitatively, animal models exhibit similar patterns of O3 net and tissue
29 dose distribution with the largest tissue dose delivered to the CAR. However, due to
30 anatomical and biochemical RT differences, the absolute value of delivered O3 dose
31 differs, with animal data obtained in resting conditions underestimating the dose to the
32 RT and presumably the resultant risk of effect for humans, especially humans during
33 exercise. Even though interspecies differences limit quantitative comparison between
34 species, the acute and chronic functional responses of laboratory animals to O3 appear
3 5 qualitatively homologous to those of the human making them a useful tool in determining
36 mechanistic and cause-effect relationships with O3 exposure.
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5.7 References
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Ahmad. S: Ahmad. A: McConville. G: Schneider. BK: Allen. CB: Manzer. R: Mason. RJ: White. CW. (2005). Lung
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AI-Hegelan. M: Tighe. RM: Castillo. C: Hollingsworth. JW. (2011). Ambient ozone and pulmonary innate
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Alexis, N: Urch, B: Tarlo, S: Corey, P: Pengelly, D: O'Byrne, P: Silverman, F. (2000). Cyclooxygenase
metabolites play a different role in ozone-induced pulmonary function decline in asthmatics compared
to normals. Inhal Toxicol 12: 1205-1224.
Alexis, N: Soukup, J: Nierkens, S: Becker, S. (2001 b). Association between airway hyperreactivity and bronchial
macrophage dysfunction in individuals with mild asthma. Am J Physiol Lung Cell Mol Physiol 280:
L369-L375.
Alexis, NE: Zhou, H: Lay, JC: Harris, B: Hernandez, ML: Lu, TS: Bromberg, PA: Diaz-Sanchez, D: Devlin, RB:
Kleeberger. SR: Peden. DB. (2009). The glutathione-S-transferase Mu 1 null genotype modulates
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6 INTEGRATED HEALTH EFFECTS OF SHORT-
TERM OZONE EXPOSURE
6.1 Introduction
1 This chapter reviews, summarizes, and integrates the evidence for various health
2 outcomes associated with short-term (i.e., hours, days, or weeks) exposures to O3.
3 Numerous controlled human exposure, epidemiologic, and toxicological studies have
4 permitted evaluation of the relationships of short-term O3 exposure with a range of
5 endpoints related to respiratory effects (Section 6.2), cardiovascular effects (Section 6.3),
6 and mortality (Sections 6.2, 6.3, and 6.6). A smaller number of studies are available to
7 assess the effects of O3 on other physiological systems such as the central nervous system
8 (Section 6.4), liver and metabolism (Section 6.5.1), and cutaneous and ocular tissues
9 (Section 6.5.2).
10 Evidence forthe major health effect categories (e.g., respiratory, cardiovascular,
11 mortality) is described in individual sections that include a brief summary of conclusions
12 from the 2006 O3 AQCD and an evaluation of recent evidence that is intended to build
13 upon evidence from previous reviews. Within each section, results are organized by
14 health endpoint (e.g., lung function, pulmonary inflammation) then by specific scientific
15 discipline (e.g., controlled human exposure, epidemiology, and toxicology). Each major
16 section (e.g., respiratory, cardiovascular, mortality) concludes with an integrated
17 summary of the findings and a conclusion regarding causality. Based upon the framework
18 described in the Preamble to this ISA, a determination of causality is made for a broad
19 health effect category, such as respiratory effects, with coherence and plausibility being
20 based on the evidence available across disciplines and also across the suite of related
21 health endpoints, including cause-specific mortality.
6.2 Respiratory Effects
22 Based on evidence integrated across human controlled exposure, epidemiologic, and
23 toxicological studies, the 2006 O3 AQCD concluded that there was clear, consistent
24 evidence of a causal relationship between short-term O3 exposure and respiratory effects
25 (U.S. EPA. 2006b). Contributing to this conclusion were consistent and coherent
26 observations across scientific disciplines of associations of short-term O3 exposures with
27 pulmonary function decrements and increases in lung inflammation, lung permeability,
28 and airway hyperresponsiveness. Collectively, these findings provided biological
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1 plausibility for associations in epidemiologic studies of short-term ambient O3 exposure
2 with respiratory symptoms and respiratory-related hospitalizations and emergency
3 department (ED) visits.
4 Controlled human exposure studies have provided strong and quantifiable exposure-
5 response data on the human health effects of O3. The most salient observations from
6 studies reviewed in the 1996 and 2006 O3 AQCDs were that: (1) young healthy adults
7 exposed to O3 concentrations > 80 ppb develop significant reversible, transient
8 decrements in pulmonary function if minute ventilation (VE) or duration of exposure is
9 increased sufficiently; (2) relative to young adults, children experience similar
10 spirometric responses but lesser symptoms from O3 exposure; (3) relative to young
11 adults, O3-induced spirometric responses are decreased in older individuals; (4) there is a
12 large degree of intersubject variability in physiologic and symptomatic responses to O3
13 but responses tend to be reproducible within a given individual over a period of
14 several months; (5) subjects exposed repeatedly to O3 for several days experience an
15 attenuation of spirometric and symptomatic responses on successive exposures, that is
16 lost after about a week without exposure; and (6) acute O3 exposure initiates an
17 inflammatory response that may persist for at least 18 to 24 hours postexposure.
18 Substantial evidence for biologically plausible O3-induced respiratory morbidity has been
19 derived from the coherence between toxicological and controlled human exposure studies
20 examining parallel endpoints. For example, O3-induced decrements in lung function have
21 also been observed in animals, and as in humans, tolerance or attenuation has been
22 demonstrated in animal models. Both humans and rodents exhibit increased airway
23 hyperresponsiveness. This is an important consequence of exposure to ambient O3,
24 because the airways are then predisposed to narrowing upon inhalation of a variety of
25 ambient stimuli. Additionally, airway hyperresponsiveness tends to resolve more slowly
26 and appears less subject to attenuation. Increased permeability and inflammation have
27 been observed in the airways of humans and animals alike after O3 exposure, although
28 these processes are not necessarily associated with immediate changes in lung function or
29 hyperresponsiveness. Furthermore, the potential relationship between repetitive bouts of
30 acute inflammation and the development of chronic respiratory disease is unknown.
31 Another feature of O3 exposure-related respiratory morbidity is impaired host defense
32 and reduced resistance to lung infection, which has been strongly supported by
33 toxicological evidence and to a limited extent by human data. Recurrent respiratory
34 infection in early life is associated with increased incidence of asthma in humans.
35 In epidemiologic studies, short-term O3-related respiratory morbidity has been assessed
36 most frequently using lung function. Several studies of healthy children attending camps
37 as well as studies of outdoor workers, groups exercising outdoors, and children with
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1 asthma support O3 effects on lung function decrements at ambient levels (U.S. EPA.
2 2006b. 1996a). In addition to lung function, ambient O3 exposure has been associated
3 with increases in respiratory symptoms (e.g., cough, wheeze, shortness of breath),
4 especially in large U.S. panel studies of children with asthma (Gent et al.. 2003;
5 Mortimer et al., 2000). The evidence across disciplines for O3 effects on a range of
6 respiratory endpoints collectively provides support for epidemiologic studies that have
7 demonstrated consistent positive associations between O3 exposure and respiratory
8 hospital admissions and ED visits, specifically during the summer or warm months. In
9 contrast with other respiratory health endpoints, the association between short-term O3
10 exposure and respiratory mortality is less clearly indicated. Although O3 has been
11 consistently associated with nonaccidental and cardiopulmonary mortality, the
12 contribution of respiratory causes to these findings has been uncertain as the few studies
13 that have examined mortality specifically from respiratory causes have reported
14 inconsistent associations with ambient O3 exposures.
15 As discussed throughout this section, consistent with the strong body of evidence
16 presented in the 2006 O3 AQCD, recent studies continue to support associations between
17 short-term O3 exposure and respiratory effects, in particular, lung function decrements in
18 controlled human exposure studies, airway inflammatory responses in toxicological
19 studies, and respiratory-related hospitalizations and ED visits. Recent epidemiologic
20 studies contribute new evidence on at-risk populations and of associations of ambient O3
21 exposures with biological markers of airway inflammation and oxidative stress, which is
22 consistent with the extensive evidence from human controlled exposure and toxicological
23 studies. Furthermore, extending the potential range of well-established O3-associated
24 respiratory effects, new multicity studies and a multicontinent study demonstrate
25 associations between short-term ambient O3 exposure and respiratory-related mortality.
6.2.1 Lung Function
6.2.1.1 Controlled Human Exposure
26 This section focuses on studies examining O3 effects on lung function and respiratory
27 symptoms in volunteers exposed, for periods of up to 8 hours to O3 concentrations
28 ranging from 40 to 500 ppb, while at rest or during exercise of varying intensity.
29 Responses to acute O3 exposures in the range of ambient concentrations include
30 decreased inspiratory capacity; mild bronchoconstriction; rapid, shallow breathing
31 patterns during exercise; and symptoms of cough and pain on deep inspiration (PDI).
32 Reflex inhibition of inspiration results in a decrease in forced vital capacity (FVC) and
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1 total lung capacity (TLC) and, in combination with mild bronchoconstriction, contributes
2 to a decrease in the forced expiratory volume in 1 second (FEVi).
3 In studies that have exposed subjects during exercise, the majority of shorter duration
4 (< 4-hour exposures) studies utilized an intermittent exercise protocol in which subjects
5 rotated between 15-minute periods of exercise and rest. A limited number of 1- to 2-hour
6 studies, mainly focusing on exercise performance, have utilized a continuous exercise
7 regime. A quasi continuous exercise protocol is common to prolonged exposure studies
8 where subjects complete 50-minute periods of exercise followed by 10-minute rest
9 periods.
10 The majority of controlled human exposure studies have been conducted within
11 chambers, although a smaller number of studies used a facemask to expose subjects to
12 O3. Little effort has been made herein to differentiate between facemask and chamber
13 exposures as FEVi and respiratory symptom responses appear minimally affected by
14 these exposure modalities. Similar responses between facemask and chamber exposures
15 have been reported for exposures to 80 and 120 ppb O3 (6.6 h, moderate quasi continuous
16 exercise, 40 L/min) and 300 ppb O3 (2 h, heavy intermittent exercise, 70 L/min) (Adams.
17 2003a, b, 2002).
18 The majority of controlled human exposure studies investigating the effects O3 are of a
19 randomized, controlled, crossover design in which subjects were exposed, without
20 knowledge of the exposure condition and in random order to clean filtered air (FA; the
21 control) and, depending on the study, to one or more O3 concentrations. The FA control
22 exposure provides an unbiased estimate of the effects of the experimental procedures on
23 the outcome(s) of interest. Comparison of responses following this FA exposure to those
24 following an O3 exposure allows for estimation of the effects of O3 itself on an outcome
25 measurement while controlling for independent effects of the experimental procedures.
26 As individuals may experience small changes in various health endpoints from exercise,
27 diurnal variation, or other effects in addition to those of O3 during the course of an
28 exposure, the term "O3-induced" is used herein to designate effects that have been
29 corrected or adjusted for such extraneous responses as measured during FA exposures.
30 Spirometry, viz., FEVi, is a common health endpoint used to assess effects of O3 on
31 respiratory health in controlled human exposure studies. In considering 6.6 hour
32 exposures to FA, group mean FEVi changes have ranged from -0.7% (McDonnell et al.,
33 1991) to 2.7% (Adams. 2006a). On average, across ten 6.6-hour exposure studies, there
34 has been a 1.0% (n=279) increase in FEVi (Kim etal.. 2011; Schelegle et al.. 2009;
35 Adams. 2006a. 2003a. 2002: Adams and Ollison. 1997: Folinsbee et al.. 1994:
36 McDonnell et al.. 1991; Horstman et al.. 1990; Folinsbee et al.. 1988). Regardless of the
37 reason for small changes in FEVi over the course of FA exposures, whether biologically
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1 based or a systematic effect of the experimental procedures, the use of FA responses as a
2 control for the assessment of responses following O3 exposure in randomized exposure
3 studies serves to eliminate alternative explanations other than those of O3 itself in causing
4 the measured responses.
5 Considering FEVi responses in young healthy adults, an O3-induced change in FEVi is
6 typically the difference between the decrement observed with O3 exposure and the
7 improvement observed with FA exposure. Noting that some healthy individuals
8 experience small improvements while others have small decrements in FEVi following
9 FA exposure, investigators have used the randomized, crossover design with each subject
10 having their own control exposure to FA to discern relatively small effects with certainty
11 since alternative explanations for these effects are controlled for by the nature of the
12 experimental design. The utility of FA control exposures becomes more apparent when
13 considering individuals with respiratory disease. The occurrence of exercise-induced
14 bronchospasm is well recognized to in patients with asthma and COPD and may be
15 experienced during both FA and O3 exposures. Absent correction for FA responses,
16 exercise-induced changes in FEVi could be mistaken for responses due to O3. This
17 biological phenomenon serves as an example to emphasize the need for a proper control
18 exposure in assessing the effects of O3 as well as the role of this control in eliminating the
19 influence of other factors on the outcomes of interest.
Pulmonary Function Effects of Ozone Exposure in Healthy Subjects
Acute Exposure of Healthy Subjects
20 The majority of controlled human exposure studies have investigated the effects of
21 exposure to O3 in young healthy nonsmoking adults (18-35 years of age). These studies
22 typically use fixed concentrations of O3 under carefully regulated environmental
23 conditions and subject activity levels. The magnitude of respiratory effects (decrements
24 in spirometry and symptomatic response) in these individuals is a function of O3
25 concentration (C), minute ventilation (VE), and exposure duration (time). Any physical
26 activity will increase minute ventilation and therefore the dose of inhaled O3. Dose of
27 inhaled O3 to the lower airways is also increased due to a shift from nasal to oronasal
28 breathing with a consequential decrease in O3 scrubbing by the upper airways. Thus, the
29 intensity of physiological response following an acute exposure will be strongly
30 associated with minute ventilation.
31 The product of C x VE x time, although actually a measure of exposure, is commonly
32 used as a surrogate for O3 dose to the respiratory tract in controlled human exposure
33 studies. The delivery of O3 to the lower respiratory tract varies as a function of breathing
34 conditions (route and pattern). And, the dose of O3 to the lower respiratory tract can vary
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1 between similarly exposed individuals. In support of the use of the product (C * VE *
2 time) as a surrogate for O3 dose, differences in FEVi responses among young healthy
3 adults (32 M, 28 F) exposed to O3 (250 ppb, 30 L/min, 2 h) do not appear to be explained
4 by intersubject differences in the fraction of inhaled O3 retained in the lung (Ultman et
5 al.. 2004). Using the product of C * VE * time as a surrogate for O3 dose is also useful in
6 distinguishing between the well defined and characterized exposure of subjects in
7 controlled human exposure studies as opposed to the use of ambient O3 concentration to
8 characterize exposure in epidemiologic studies.
9 For healthy young adults exposed at rest for 2 hours, 500 ppb is the lowest O3
10 concentration reported to produce a statistically significant O3-induced group mean FEVi
11 decrement of 6.4% (n=10) (Folinsbee et al.. 1978) to 6.7% (n=13) (Horvath et al.. 1979).
12 Airway resistance was not clearly affected during at-rest exposure to these
13 O3 concentrations. When exposed to 200 ppb for 2.25 h during intermittent periods of rest
14 and brisk walking, young healthy subjects (83 M, 55 F) show a statistically significant
15 group mean FEVi decrement of 8.8% following O3 exposure (Que etal.). For exposures
16 of 1-2 hours to > 120 ppb O3, statistically significant symptomatic responses and effects
17 on FEVi are observed when VE is sufficiently increased by exercise (McDonnell et al..
18 1999). For instance, 5% of young healthy adults exposed to 400 ppb for 2 h during rest
19 experienced pain on deep inspiration. Respiratory symptoms were not observed at lower
20 exposure concentrations (120-300 ppb) or with only 1 h of exposure. However, when
21 exposed to 120 ppb for 2 h during moderate intermittent exercise, 9% of individuals
22 experienced pain on deep inspiration, 5% experienced cough, and 4% experienced
23 shortness of breath. With very heavy continuous exercise (VE = 89 L/min), an O3-induced
24 group mean decrement of 9.7% in FEVi has been reported for healthy young adults
25 exposed for 1 hour to 120 ppb O3 (Gong et al., 1986). Symptoms are present and
26 decrements in forced expiratory volumes and flows occur at 160-240 ppb O3 following 1
27 hour of continuous heavy exercise (VE « 55 to 90 L/min (Gong et al., 1986; Avol et al..
28 1984; Folinsbee et al.. 1984; Adams and Schelegle. 1983) and following 2 hours of
29 intermittent heavy exercise (VE « 65-68 L/min) (Linn etal.. 1986; Kulleetal.. 1985;
30 McDonnell et al.. 1983). With heavy intermittent exercise (15-min intervals of rest and
31 exercise [VE = 68 L/min]), symptoms of breathing discomfort and a group mean O3-
32 induced decrement of 3.4% in FEVi occurred in young healthy adults exposed for 2
33 hours to 120 ppb O3 (McDonnell et al.. 1983).'
1 In total, subjects were exposed to O3 for 2.5 hours. Intermittent exercise periods, however, were only conducted for the first 2
hours of exposure and FENA, was determined 5 minutes after the exercise was completed.
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20 -
•D 5^
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i «
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n Horstman etal. (1990)
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0.02 0.04 0.06 0.08 0.1
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0.12
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Studies appearing in the figure legends are: Adams (2006a, 2003a. 2002. 1998). Folinsbee et al. (1988). Horstman et al. (1990).
Kim et al. (2011), McDonnell et al. (2007: 1991), and Schelegle et al. (2009).
Top, panel A: all studies exposed subjects to a constant (square-wave) concentration in a chamber, except Adams (1998) where a
facemask was used. The McDonnell et al. (2007) curve illustrates the predicted FENA, decrement at 6.6 hours as a function of ozone
concentration for a 23-year old (the average age of subjects that participated in the illustrated studies). Note that this curve was not
"fitted" to the plotted data. Error bars (where available) are the standard error of responses. Bottom, panel B: all studies used
constant (square-wave) exposures in a chamber unless designated as triangular (t) and/or facemask (m) exposures.
Figure 6-1 Cross-study comparison of mean ozone-induced FEVi decrements
following 6.6 hours of exposure to ozone. During each hour of the
exposures, subjects were engaged in moderate quasi continuous
exercise (40 L/min) for 50 minutes and rest for 10 minutes. Following the
third hour, subjects had an additional 35-minute rest period for lunch. The
data at 0.06, 0.08 and 0.12 ppm have been offset for illustrative purposes.
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6-7
September 2011
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1 For prolonged (6.6 hours) exposures relative to shorter exposures, significant pulmonary
2 function responses and symptoms have been observed at lower O3 concentrations and at a
3 moderate level of exercise (VE = 40 L/min). The results from studies using 6.6 hours of
4 constant or square-wave (S-W) exposures to between 40 and 120 ppb are illustrated in
5 Figure 6-l(A). Figure 6-l(B) focuses on the range from 40 to 80 ppb and includes
6 triangular exposure protocols as well as facemask exposures. Exposure to 40 ppb for 6.6
7 hours produces small, statistically insignificant changes in FEVi that are relatively
8 similar to responses from FA exposure (Adams. 2002). Volunteers exposed to 60 ppb O3
9 experience group mean O3-induced FEVi decrements of about 3% (Kimetal.. 2011;
10 Brown et al., 2008) (Adams. 2006a)'; those exposed to 80 ppb have group mean
11 decrements which range from 6 to 8% (Adams. 2006a. 2003a: McDonnell et al.. 1991;
12 Horstman et al., 1990); at 100 ppb, group mean decrements range from 8 to 14%
13 (McDonnell et al.. 1991; Horstman et al.. 1990): and at 120 ppb, group mean decrements
14 of 13 to 16% are observed (Adams. 2002; Horstman et al.. 1990; Folinsbee et al.. 1988).
15 As illustrated in Figure 6-1, there is a smooth dose-response curve without evidence of a
16 threshold for exposures between 40 and 120 ppb O3. Taken together, these data indicate
17 that mean FEVi is clearly decreased by 6.6-h exposures to 60 ppb O3 and higher
18 concentrations in subjects performing moderate exercise.
19 As opposed to constant or S-W concentration patterns used in the studies described
20 above, many studies conducted at the levels of 40-80 ppb have used variable O3
21 concentration patterns. It has been suggested that a triangular (variable concentration)
22 exposure profile can potentially lead to higher FEVi responses than S-W profiles despite
23 having at the same average O3 concentration over the exposure period. Hazucha et al.
24 (1992) were the first to investigate the effects of variable versus constant concentration
25 exposures on responsiveness to O3. In their study, volunteers were randomly exposed to a
26 triangular concentration profile (averaging 120 ppb over the 8-h exposure) that increased
27 linearly from 0-240 ppb for the first 4 hours of the 8-h exposure, then decreased linearly
28 from 240 to 0 ppb over the next 4 hours of the 8-h exposure, and to an S-W exposure of
29 120 ppb O3 for 8 hours. While the total inhaled O3 doses at 4 hours and 8 hours for the S-
30 W and the triangular concentration profile were almost identical, the FEVi response was
31 dissimilar. For the S-W exposure, FEVi declined ~5% by the fifth hour and then
32 remained at that level. With the triangular O3 profile, there was minimal FEVi response
33 over the first 3 hours followed by a rapid decrease in FEVi (-10.3%) over the next 3
1 Adams (2006a) did not find effects on FE\A at 60 ppb to be statistically significant. In an analysis of the Adams (2006a) data,
even after removal of potential outliers, Brown et al. (2008) found the average effect on FENA, at 60 ppb to be small, but highly
statistically significant (p < 0.002) using several common statistical tests.
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1 hours. During the seventh and eighth hours, mean FEVi decrements improved to -6.3%
2 as the O3 concentration decreased from 120 to 0 ppb (mean = 60 ppb). These findings
3 illustrate that the severity of symptoms and the magnitude of spirometric responses are
4 time-dependent functions of O3 delivery rate with periods of both effect development and
5 recovery during the course of an exposure.
6 Subsequently, others have also demonstrated that variable concentration exposures can
7 elicit greater FEVi and symptomatic responses than do S-W exposures (Adams. 2006a. b,
8 2003a). Adams (2006b) reproduced the findings of Hazucha et al. (1992) at 120 ppb.
9 However, Adams (2006a, 2003a) found that responses from an 80 ppb O3 (average)
10 triangular exposure did not differ significantly from those observed in the 80 ppb O3 S-W
11 exposure at 6.6 hours. Nevertheless, FEVi and symptoms were significantly different
12 from pre-exposure at 4.6 hours (when the O3 concentration was 150 ppb) in the triangular
13 exposure, but not until 6.6 hours in the S-W exposure. At the lower O3 concentration of
14 60 ppb, no temporal pattern differences in FEVi responses between S-W and triangular
15 exposure profiles could be discerned (Adams. 2006a). However, total symptom scores
16 were significantly increased for the 60 ppb triangular (but not the S-W) exposure
17 following 5.6 and 6.6 hours of exposure. At 80 ppb, respiratory symptoms tended to
18 increase more rapidly during the triangular than S-W exposure protocol, but then
19 decreased during the last hour of exposure to be less for the triangular than the S-W
20 exposure at 6.6 h. Both total symptom scores and pain on deep inspiration were
21 significantly increased following exposures to 80 ppb relative to all other exposure
22 protocols, i.e., FA, 40, and 60 ppb exposures. Following the 6.6-hour exposures,
23 respiratory symptoms at 80 ppb were rougly 2-3 times greater than observed at 60 ppb.
24 At 40 ppb, triangular and S-W patterns produced spirometric and subjective symptom
25 responses similar to FA exposure (Adams. 2006a. 2002).
26 For exposures of 60 ppb and greater, these studies (Adams. 2006a. b, 2003a; Hazucha et
27 al.. 1992) demonstrate that during triangular exposure protocols, volunteers exposed
28 during moderate exercise (VE = 40 L/min) may develop greater spirometric and/or
29 symptomatic responses during and following peak O3 concentrations as compared to
30 responses over the same time interval of S-W exposures. This observation is not
31 unexpected since the inhaled dose rate during peaks of the triangular protocols
32 approached twice that of the S-W protocols, e.g., 150 ppb versus 80 ppb peak
33 concentration. At time intervals toward the end of an exposure, O3 delivery rates for the
34 triangular protocols were less than those of S-W. At these later time intervals, there is
35 some recovery of responses during triangular exposure protocols, whereas there is a
36 continued development of or a plateau of responses in the S-W exposure protocols. Thus,
37 responses during triangular protocols relative to S-W protocols may be expected to
38 diverge and be greater following peak exposures and then converge toward the end of an
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1 exposure. The ensuing discussion on exposures between 40 and 80 ppb will focus on
2 postexposure effects where the influence of triangular and S-W concentration patterns are
3 minimal, i.e., FEVi pre-to-post effects are similar (although not identical) between
4 triangular and S-W protocols having equivalent average exposure concentrations.
5 Schelegle et al. (2009) recently investigated the effects of 6.6 hours variable O3 exposure
6 protocols at mean concentrations of 60, 70, 80, and 87 ppb on respiratory symptoms and
7 pulmonary function in young healthy adults (16 F, 15 M; 21.4 ± 0.6 years) exposed
8 during moderate quasi continuous exercise (VE = 40 L/min). The mean FEVi (±standard
9 error) decrements at 6.6 hours (end of exposure relative to pre-exposure) were -0.80 ±
10 0.90%, 2.72 ± 1.48%, 5.34 ± 1.42%, 7.02 ± 1.60%, and 11.42 ± 2.20% for exposure to
11 FA, 60, 70, 80, and 87 ppb O3, respectively. Statistically significant decrements in FEVi
12 and increases in total subjective symptom scores (p < 0.05) were found following
13 exposure to mean concentrations of 70, 80, and 87 ppb O3 relative to FA. Statistically
14 significant effects were not found at 60 ppb. One of the expressed purposes of the
15 Schelegle et al. (2009) study was to determine the minimal mean O3 concentration that
16 produces a statistically significant decrement in FEVi and symptoms in healthy
17 individuals completing 6.6-h exposure protocols. At 70 ppb, Schelegle et al. (2009)
18 observed a statistically significant O3-induced of 6.1%. At 60 ppb, an O3-induced 3.5%
19 FEVi decrement was not found to be statistically significant. However, this effect is
20 similar in magnitude to the 2.9% FEVi decrement at 60 ppb observed by Adams (2006a)
21 that was found to be statistically significant by Brown et al. (2008).
22 More recently, Kim et al. (2011) investigated the effects of a 6.6-h exposure to 60 ppb O3
23 during moderate quasi continuous exercise (VE = 40 L/min) on pulmonary function and
24 respiratory symptoms in young healthy adults (32 F, 27 M; 25.0 ± 0.5 year) that were
25 roughly half GSTM1-null and half GSTM1-positive. Sputum neutrophil levels were also
26 measured in a subset of the subjects (13 F, 11 M). The mean FEVi (±standard error)
27 decrements at 6.6 hours (end of exposure relative to pre-exposure) were significantly
28 different (p = 0.008) between the FA (0.002 ± 0.46%) and O3 (1.76 ± 0.50%) exposures.
29 The inflammatory response following O3 exposure was also significantly (p<0.001)
30 increased relative to the FA exposure. Respiratory symptoms were not affected by O3
31 exposure. There was also no significant effect of GSTM1 genotype on FEVi or
32 inflammatory responses.
33 Consideration of the minimal O3 concentration producing statistically significant effects
34 on FEVi following 6.6-h exposures warrants additional discussion. As discussed above,
35 numerous studies have demonstrated statistically significant O3-induced group mean
36 FEVi decrements of 6-8% at 80 ppb. Schelegle et al. (2009) have now reported
37 statistically significant O3-induced group mean FEVi decrement of 6%, as well as
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1 respiratory symptoms, at 70 ppb. At 60 ppb, there is information available from 4
2 separate studies (Adams. 1998)1 (Kim etal.. 2011: Schelegle et al.. 2009: Adams. 2006a).
3 The group mean O3-induced FEVi decrements observed in these studies were 3.6% by
4 Adams (1998)2, 2.8% (triangular exposure) and 2.9% (S-W exposure) by Adams (2006a).
5 3.5% by Schelegle et al. (2009). and 1.8% by Kim et al. (2011). Based on data from these
6 four studies, at 60 ppb, the weighted-average group mean O3-induced FEVi decrement
7 (i.e., adjusted for FA responses) is 2.7% (n=150) (Kim etal.. 2011: Schelegle et al.. 2009:
8 Adams. 2006a. 1998). Although not consistently statistically significant, these group
9 mean changes in FEVi at 60 ppb are consistent between studies, i.e., none observed an
10 average improvement in lung function following a 6.6-h exposure to 60 ppb O3. Indeed,
11 as was illustrated in Figure 6-1, the FEVi responses at 60 ppb fall on a smooth dose-
12 response curve for exposures between 40 and 120 ppb O3. Furthermore, in a re-analysis
13 of the 60 ppb S-W data from Adams (2006a). Brown et al. (2008) found the mean effects
14 on FEVi to be highly statistically significant (p<0.002) using several common statistical
15 tests even after removal of 3 potential outliers. The time-course and magnitude of FEVi
16 responses at 40 ppb resemble those occurring during FA exposures (Adams. 2006a.
17 2002). Taken together, the available evidence shows that detectable effects of O3 on
18 group mean FEVi persist down to 60 ppb, but not 40 ppb in young healthy adults
19 exposed for 6.6 hours during moderate exercise.
20 In addition to overt effects of O3 exposure on the large airways indicated by spirometric
21 responses, O3 exposure also affects the function of the small airways and parenchymal
22 lung. Foster et al. (1997: 1993) examined the effect of O3 on ventilation distribution. In
23 healthy adult males (n=6; 26.7 ± 7 years old) exposed to O3 (330 ppb with light
24 intermittent exercise for 2 h), there was a significant reduction in ventilation to the lower
25 lung (31% of lung volume) and significant increases in ventilation to the upper- and
26 middle-lung regions (Foster et al.. 1993). In a subsequent study of healthy males (n=15;
27 25.4 ± 2 years old) exposed to O3 (350 ppb with moderate intermittent exercise for 2.2 h),
28 O3 exposure caused a delayed gas washout (Foster etal.. 1997). The pronounced slow
29 phase of gas washout following O3 exposure represented a 24% decrease in the washout
30 rate. A day following O3 exposure, 50% of the subjects still had (or developed) a delayed
31 washout relative to the pre- O3 maneuver. These studies suggest a prolonged O3 effect on
32 the small airways and ventilation distribution in healthy young individuals.
1 The American Petroleum Institute has declined to provide a copy of this report to EPA.
2 This information is from page 133 of Adams (2006a). This decrement may be increased due to a target VE of 23 L/min/m2 BSA
relative to other studies with which it is listed having the target VE of 20 L/min/m2 BSA. It should also be noted that subjects were
exposed via a facemask in this study. However, Adams (2003a, b, 2002) found very similar FE\A responses between facemask and
chamber exposures.
Draft - Do Not Cite or Quote 6-11 September 2011
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1 There is a rapid recovery of O3-induced spirometric responses and symptoms; 40 to 65%
2 recovery appears to occur within about 2 hours following exposure (Tolinsbee and
3 Hazucha. 1989). For example, following a 2-h exposure to 400 ppb O3 with intermittent
4 exercise, Nightingale et al. (2000) observed a 13.5% mean decrement in FEVi. By 3
5 hours postexposure, however, only a 2.7% FEVi decrement persisted. Partial recovery
6 also occurs following cessation of exercise despite continued exposure to O3 (Tolinsbee
7 etal.. 1977) and at low O3 concentrations during exposure (Hazucha etal.. 1992). A
8 slower recovery phase, especially after exposure to higher O3 concentrations, may take at
9 least 24 hours to complete (Folinsbee and Hazucha, 2000; Folinsbee et al., 1993).
10 Repeated daily exposure studies at higher concentrations typically show that FEVi
11 response to O3 is enhanced on the second day of exposure. This enhanced response
12 suggests a residual effect of the previous exposure, about 22 hours earlier, even though
13 the pre-exposure spirometry may be the same as on the previous day. The absence of the
14 enhanced response with repeated exposure at lower O3 concentrations may be the result
15 of a more complete recovery or less damage to pulmonary tissues (Folinsbee etal.. 1994).
Intersubject Variability in Response of Healthy Subjects
16 Consideration of group mean changes is important in discerning if observed effects are
17 due to O3 exposure rather than chance alone. Inter-individual variability in responses is,
18 however, considerable and pertinent to assessing the fraction of the population that might
19 actually be affected during an O3 exposure. Hackney et al. (1975) first recognized a wide
20 range in the sensitivity of subjects to O3. The range in the subjects' ages (29 to 49 years)
21 and smoking status (0 to 50 pack years) in the Hackney et al. (1975) study are now
22 understood to affect the spirometric and symptomatic responses to O3. Subsequently,
23 DeLucia and Adams (1977) examined responses to O3 in six healthy non-smokers and
24 found that two exhibited notably greater sensitivity to O3. Since that time, numerous
25 studies have documented considerable variability in responsiveness to O3 even in subjects
26 recruited to assure homogeneity in factors recognized or presumed to affect responses.
27 An individual's FEVi response to a 2-h O3 exposure is generally reproducible over
28 several months and presumably reflects the intrinsic responsiveness of the individual to
29 O3 (Hazucha et al., 2003; McDonnell et al., 1985a). The frequency distribution of
30 individual FEVi responses following these relatively short exposures becomes skewed as
31 the group mean response increases, with some individuals experiencing large reductions
32 in FEVi (Weinmann et al.. 1995c: Kulle et al.. 1985). For 2-h exposures with intermittent
33 exercise causing a predicted average FEVi decrement of 10%, individual decrements
34 ranged from approximately 0 to 40% in white males aged 18-36 years (McDonnell et al..
35 1997). For an average FEVi decrement of 13%, Ultman et al. (2004) reported FEVi
36 responses ranging from a 4% improvement to a 56% decrement in young healthy adults
Draft - Do Not Cite or Quote 6-12 September 2011
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1
2
(32 M, 28 F) exposed for 1 hour to 250 ppb O3. One-third of the subjects had FEVi
decrements of >15%, and 7% of the subjects had decrements of >40%.
4
5
6
7
8
9
10
11
12
13
14
15
16
M 35 |
O 3Q .
5" OT
3
CO 20-
0 15-
M 10-
o
Oi 5 '
Q- n.
n
, — .
. .
Oppb
0%
i — i
-
•
.
nl~l
. .
60 ppb
16%
^ n
n
__
. .
70 ppb
i—i
19%
Hn n
nn
80 ppb
I — i
29%
In] In
10 20 30 -10 0
FEV, Decrement (%)
• 10 0 10 20 30
Source: Adapted with permission of American Thoracic Society (Schelegle et al., 2009)
During each hour of the exposures, subjects were engaged in moderate quasi continuous exercise (40 L/min) for 50 minutes and
rest for 10 minutes. Following the third hour, subjects had an additional 35 minute rest period for lunch. Subjects were exposed to a
triangular ozone concentration profile having the average ozone concentration provided in each panel. As average ozone
concentration increased, the distribution of responses became asymmetric with a few individuals exhibiting large FEVi decrements.
The percentage indicated in each panel is the portion of subjects having a FEVi decrement in excess of 10%.
Figure 6-2 Frequency distributions of FEVi decrements observed by Schelegle
et al. (2009) in young healthy adults (16 F, 15 M) following 6.6-h
exposures to ozone or filtered air.
Consistent with the 1- to 2-h studies, the distribution of individual responses following
6.6-h exposure studies becomes skewed with increasing exposure concentration and
magnitude of the group mean FEVi response (McDonnell. 1996). Figure 6-2 illustrates
frequency distributions of individual FEVi responses observed in 31 young healthy adults
following 6.6-h exposures between 0 and 80 ppb. Schelegle et al. (2009) found >10%
FEVi decrements in 16, 19, 29, and 42% of individuals exposed for 6.6 hours to 60, 70,
80, and 87 ppb, respectively. Just as there are differences in mean decrements between
studies having similar exposure scenarios (Figure 6-1 at 80 and 120 ppb), there are also
differences in the proportion of individuals affected with >10% FEVi decrements. At
80 ppb, the proportion affected with >10% FEVi decrements was 17% (n=30) by Adams
(2006a)'. 26% (n=60) by McDonnell (1996). and 29%(n=31) by Schelegle et al. (2009).
At 60 ppb, the proportion with >10% FEVi decrements was 20% (n=30) by Adams
(1998)2. 3% (n=30) by Adams (2006a)5, 16% (n=31) by Schelegle et al. (2009). and 5%
(n=59) by Kim et al. (2011). Based on these studies, the weighted average proportion of
1 Not assessed by Adams (2006a), the proportion was provided in Figure 8-1B of U.S. EPA (2006b).
2 This information is from page 761 of Adams (2002).
Draft - Do Not Cite or Quote
6-13
September 2011
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1 individuals with >10% FEVi decrements is 10% following exposure to 60 ppb. Due to
2 limited data within the published papers, these proportions were not corrected for
3 responses to FA exposure where lung function typically improves in healthy adults. For
4 example, uncorrected versus O3-induced (i.e., adjusted for response during FA exposure)
5 proportions of individuals having >10% FEVi decrements in the Adams (2006a)1 study
6 were, respectively, 3% versus 7% at 60 ppb and 17% versus 23% at 80 ppb. Thus,
7 uncorrected proportions underestimate the actual fraction of healthy individuals affected.
8 Given considerable inter-individual variability in responses, the interpretation of
9 biologically small group mean decrements requires careful consideration. Following
10 prolonged 6.6-h exposures to an average level of 60 ppb O3, data available from four
11 studies yield a weighted-average group mean O3-induced FEVi decrement (i.e., adjusted
12 for FA responses) of 2.7% (n=150) (Kim etal.. 2011; Schelegle et al.. 2009; Adams.
13 2006a. 1998). The data from these studies also yield a weighted-average proportion
14 (uncorrected for FA responses) of subjects with >10% FEVi decrements of 10% (n=150)
15 (Kim etal.. 2011: Schelegle et al.. 2009: Adams. 2006a. 1998). In an individual with
16 relatively "normal" lung function, recognizing technical and biological variability in
17 measurements, confidence can be given that within-day changes in FEVi of > 5% are
18 clinically meaningful (Pellegrino et al.. 2005: ATS. 1991). Here focus is given to
19 individuals with >10% decrements in FEVi since some individuals in the Schelegle et al.
20 (2009) study experienced 5-10% FEVi decrements following exposure to FA. A 10%
21 FEVi decrement is also generally accepted as an abnormal response and as reasonable
22 criterion for assessing exercise-induced bronchoconstriction (Dryden et al.. 2010: ATS.
23 2000a). The data are not available in the published papers to determine the O3-induced
24 proportion for either the Adams (1998) or Schelegle et al. (2009) studies. As already
25 stated, however, this uncorrected proportion likely underestimates that actual proportion
26 of healthy individuals experiencing O3-induced FEVi decrements in excess of 10%.
27 Therefore, by considering uncorrected responses and those individuals having >10%
28 decrements, 10% is an underestimate of the proportion of healthy individuals that are
29 likely to experience clinically meaningful changes in lung function following exposure
30 for 6.6 hours to 60 ppb O3 during moderate exercise. Of the studies conducted at 60 ppb,
31 only Kim et al. (2011) reported FEVi decrements at 60 ppb to be statistically significant.
32 Although, Brown et al. (2008) found those from Adams (2006a) to be highly statistically
33 significant. Though group mean decrements are biologically small and generally do not
34 attain statistical significance, a considerable fraction of exposed individuals experience
35 clinically meaningful decrements in lung function.
1 Not assessed by Adams (2006a), uncorrected and OS-induced proportions are from Figures 8-1B and 8-2, respectively, of the
2006O3AQCD(2006b).
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Responses in Individuals with Pre-Existing Disease
1 Individuals with respiratory disease are of primary concern in evaluating the health
2 effects of O3 because a given change in function is likely to have more impact on a
3 person with preexisting function impairment and reduced reserve.
4 Possibly due to the age of subjects studied, patients with COPD performing light to
5 moderate exercise do not generally experience statistically significant pulmonary
6 function decrements following 1- and 2-h exposures to < 300 ppb O3 (Kehrl et al.. 1985;
7 Linn et al.. 1983; Linn et al.. 1982b: Solic et al.. 1982). Following a 4-hour exposure to
8 240 ppb O3 during exercise, Gong et al. (1997b) found an O3-induced FEVi decrement of
9 8% in COPD patients which was not statistically different from the decrement of 3% in
10 healthy subjects. Demonstrating the need for control exposures and presumably due to
11 exercise, four of the patients in the Gong et al. (1997b) study had FEVi decrements of
12 >14% following both the FA and O3 exposures. Although the clinical significance is
13 uncertain, small transient decreases in arterial blood oxygen saturation have also been
14 observed in some of these studies.
15 Based on studies reviewed in the 1996 and 2006 O3 AQCDs, asthmatic subjects appear to
16 be at least as sensitive to acute effects of O3 as healthy nonasthmatic subjects. Horstman
17 et al. (1995) found the O3-induced FEVi decrement in mild-to-moderate asthmatics to be
18 significantly larger than in healthy subjects (19% versus 10%, respectively) exposed to
19 160 ppb O3 during exercise for 7.6-h exposure. In asthmatics, a significant positive
20 correlation between O3-induced spirometric responses and baseline lung function was
21 observed, i.e., responses increased with severity of disease. Such differences in
22 pulmonary function between asthmatics and healthy individuals were not found in shorter
23 duration studies. Alexis et al. (2000) and Torres et al. (1996) reported a tendency for
24 slightly greater FEVi decrements in asthmatics than healthy subjects. Several studies
25 reported similar responses between asthmatics and healthy individuals (Scannell et al..
26 1996; Hiltermann et al.. 1995; Bashaetal.. 1994). The lack of differences in the
27 Hiltermann et al. (1995) and Basha et al. (1994) studies was not surprising, however,
28 given extremely small sample sizes and corresponding lack of statistical power. One
29 study reported a tendency for asthmatics to have smaller O3-induced FEVi decrements
30 than healthy subjects (3% versus 8%, respectively) when exposed to 200 ppb O3 for 2
31 hours during exercise (Mudway et al.. 2001). However, the asthmatics in that study also
32 tended to be older than the healthy subjects, which could partially explain their lesser
33 response since FEVi responses to O3 diminish with age.
34 Some, but not all, studies have also reported that asthmatics have a somewhat
35 exaggerated airway inflammatory response to acute O3 exposure relative to healthy
36 control subjects (Holz et al.. 2002; Peden. 2001; Newson et al.. 2000; Hiltermann et al..
Draft - Do Not Cite or Quote 6-15 September 2011
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1 1999: Michelson et al.. 1999; Vagaggini et al.. 1999; Hiltermann et al.. 1997; Peden et
2 al.. 1997: Scannell et al.. 1996: Peden et al.. 1995: Bashaetal.. 1994: McBride et al..
3 1994). For example, at 18 hours post-O3 exposure (200 ppb, 4 hours with exercise) and
4 corrected for FA responses, Scannell et al. (1996) found significantly increased
5 neutrophils in 18 asthmatics (12%) compared to 20 healthy subjects (4.5%). This
6 difference in inflammatory response was observed despite no group differences in
7 spirometric responses to O3.
8 Vagaggini et al. (2010) exposed mild-to-moderate asthmatics (n=23; 33 ± 11 years) to
9 300 ppb O3 for 2 hours with moderate exercise. Although the group mean O3-induced
10 FEVi decrement was only 4%, eight subjects were categorized as "responders" with
11 >10% FEVi decrements. There were no baseline differences between responders and
12 nonresponders. At 6 hours post O3 exposure, sputum neutrophils were significantly
13 increased by 15% relative to FA in responders. The neutrophil increase in responders was
14 also significantly greater than the 0.2% increase in nonresponders. Across all subjects,
15 there was a significant (r=0.61, p = 0.015) correlation between changes in FEVi and
16 changes in sputum neutrophils. Prior studies have reported that inflammatory responses
17 do not appear to be correlated with lung function responses in either asthmatic or healthy
18 subjects (Holzetal.. 1999: Balmes et al.. 1997: Balmes et al.. 1996: Devlin et al.. 1991).
19 Interestingly, the nonresponders in the Vagaggini et al. (2010) study experienced a
20 significant O3-induced 11.3% increase in sputum eosinophils, while responders had an
21 nonsignificant 2.6% decrease. Six of the subjects were NQO1 wild type and GSTM1 null,
22 but this genotype was not found to be associated with the changes in lung function or
23 inflammatory responses to O3.
24 A few recent studies have evaluated the effects of corticosteroid usage on the response of
25 asthmatics to O3. Vagaggini et al. (2007) evaluated whether corticosteroid usage would
26 prevent O3-induced lung function decrements and inflammatory responses in a group of
27 subjects with mild persistent asthma (n=9; 25 ± 7 years). In this study, asthmatics were
28 randomly exposed on four occasions to 270 ppb O3 or FA for 2 hours with moderate
29 exercise. Exposures were preceded by four days of treatment with prednisone or placebo.
30 Pretreatment with corticosteroids prevented an inflammatory response in induced sputum
31 at 6 hours postexposure. FEVi responses were, however, not prevented by corticosteroid
32 treatment and were roughly equivalent to those observed following placebo. Vagaggini et
33 al. (2001) also found budesonide to decrease airway neutrophil influx in asthmatics
34 following O3 exposure. In contrast, inhalation of corticosteroid budesonide failed to
35 prevent or attenuate O3-induced responses in healthy subjects as assessed by
36 measurements of lung function, bronchial reactivity and airway inflammation
37 (Nightingale et al.. 2000). High doses of inhaled fluticasone and oral prednisolone have
Draft - Do Not Cite or Quote 6-16 September 2011
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1 each been reported to reduce inflammatory responses to O3 in healthy individuals (Holz
2 etal.. 2005).
3 More recently, Stenfors et al. (2010) exposed persistent asthmatics (n=13; aged 33 years)
4 receiving chronic inhaled corticosteroid therapy to 200 ppb O3 for 2 hours with moderate
5 exercise. An average O3-induced FEVi decrement of 8.4% was observed, whereas, only a
6 3.0% FEVi decrement is predicted for similarly exposed age-matched healthy controls
7 (McDonnell et al.. 2007). At 18 hours postexposure, there was a significant O3-induced
8 increase in bronchioalveolar lavage (BAL) neutrophils, but not eosinophils. Bronchial
9 biopsy also showed a significant O3-induced increase in mast cells. This study suggests
10 that the protective effect of acute corticosteroid therapy against inflammatory responses
11 to O3 in asthmatics demonstrated by Vagaggini et al. (2007) may be lost with continued
12 treatment regime s.
Factors Modifying Responsiveness to Ozone
13 Physical activity increases VE and therefore the dose of inhaled O3. Consequently, the
14 intensity of physiological response during and following an acute O3 exposure will be
15 strongly associated with minute ventilation. Apart from inhaled O3 dose and related
16 environmental factors (e.g., repeated daily exposures), individual-level factors, such as
17 health status, age, gender, ethnicity, race, smoking habit, diet, and socioeconomic status
18 (SES) have been considered as potential modulators of a physiologic response to such
19 exposures.
20 Children, adolescents, and young adults (<18 years of age) appear, on average, to have
21 nearly equivalent spirometric responses to O3, but have greater responses than middle-
22 aged and older adults when exposed to comparable O3 doses (U.S. EPA, 1996a).
23 Symptomatic responses to O3 exposure, however, appear to increase with age until early
24 adulthood and then gradually decrease with increasing age (U.S. EPA. 1996a). For
25 example, healthy children (aged 8-11 y) exposed to 120 ppb O3 (2.5 h; heavy intermittent
26 exercise) experienced similar spirometric responses but lesser symptoms than similarly
27 exposed young healthy adults (McDonnell et al.. 1985b). For subjects aged 18-36 years,
28 McDonnell et al. (1999) reported that symptom responses from O3 exposure also
29 decrease with increasing age. Diminished symptomatic responses in children and the
30 elderly might put these groups at increased risk for continued O3 exposure, i.e., a lack of
31 symptoms may result in their not avoiding or ceasing exposure. Once lung growth and
32 development reaches the peak (18-20 years of age in females and early twenties in
33 males), pulmonary function, which is at its maximum as well, begins to decline
34 progressively with age as does O3 sensitivity.
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1 In healthy individuals, the fastest rate of decline in O3 responsiveness appears between
2 the ages of 18 and 35 years (Passannante et al.. 1998; Seal et al.. 1996). more so for
3 females then males (Hazucha et al., 2003). During the middle age period (35-55 years),
4 O3 sensitivity continues to decline but at a much lower rate. Beyond this age (>55 years),
5 acute O3 exposure elicits minimal spirometric changes. Whether the same age-dependent
6 pattern of O3 sensitivity decline also holds for nonspirometric pulmonary function,
7 airway reactivity or inflammatory endpoints has not been determined. Although there is
8 considerable evidence that spirometric and symptomatic responses to O3 exposure
9 decrease with age beyond young adulthood, this evidence comes from cross-sectional
10 analyses and has not been confirmed by longitudinal studies of the same individuals.
11 Several studies have suggested that physiological differences between sexes may
12 predispose females to a greater susceptibility to O3. In females, lower plasma and nasal
13 lavage fluid (NLF) levels of uric acid (the most prevalent antioxidant), the initial defense
14 mechanism of O3 neutralization in airway surface liquid, may be a contributing factor
15 (Housley et al.. 1996). Consequently, reduced absorption of O3 in the upper airways may
16 promote its deeper penetration. Dosimetric measurements have shown that the absorption
17 distribution of O3 is independent of gender when absorption is normalized to anatomical
18 dead space (Bushetal.. 1996). Thus, a gender-related differential removal of O3 by uric
19 acid seems to be minimal. In general, the physiologic response of young healthy females
20 to O3 exposure appears comparable to the response of young males (Hazucha et al..
21 2003). Several studies have investigated the effects of the menstrual cycle on responses to
22 O3 in healthy young women. In a study of 9 women exposed during exercise to 300 ppb
23 O3 for an hour, Fox et al. (1993) found lung function responses to O3 significantly
24 enhanced during the follicular phase relative to the luteal phase. However, Weinmann et
25 al. (1995a) found no difference in responses between the follicular and luteal phases as
26 well as no significant differences between 12 males and 12 females exposed during
27 exercise to 350 ppb O3 for 2.15 h. Seal et al. (1996) also reported no effect of menstrual
28 cycle phase in their analysis of responses of 150 women (n=25 per exposure group; 0,
29 120, 240, 300, and 400 ppb O3). Seal et al. (1996) conceded that the methods used by Fox
30 et al. (1993) more precisely defined menstrual cycle phase.
31 Only two controlled human exposure studies have assessed differences in lung function
32 responses between races. Seal et al. (1993) compared lung function responses of whites
33 (93 M, 94 F) and blacks (undefined ancestry; 92 M, 93 F) exposed to a range of O3
34 concentrations (0-400 ppb). The main effects of gender-race group and O3 concentration
35 were statistically significant (both at p < 0.001), although the interaction between gender-
36 race group and O3 concentration was not significant (p = 0.13). These findings indicate
37 some overall difference between the gender-race groups that is independent of O3
38 concentration, i.e., the concentration-response curves for the four gender-race groups are
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1 parallel. In a multiple comparison procedure on data collapsed across all O3
2 concentrations for each gender-race group, both black men and black women had
3 significantly larger decrements in FEVi than did white men. The authors noted that the
4 O3 dose per unit of lung tissue would be greater in blacks and females than whites and
5 males, respectively. That this difference in tissue dose might have affected responses to
6 O3 cannot be ruled out. The college students recruited for the Seal et al. (1993) study are
7 probably from better educated and SES advantaged families, thus reducing potential
8 influence of these variables on results. In a follow-up analysis, Seal et al. (1996) reported
9 that, of three SES categories, individuals in the middle SES category showed greater
10 concentration-dependent decline in percent-predicted FEVi (4-5% at 400 ppb O3) than
11 low and high SES groups. The authors did not have an "immediately clear" explanation
12 for this finding.
13 More recently, Que et al. assessed pulmonary responses in blacks of African American
14 ancestry (22 M, 24 F) and Caucasians (55 M, 28 F) exposed to 220 ppb O3 for 2.25 h
15 (alternating 15 min periods of rest and brisk treadmill walking). On average, the black
16 males experienced a 16.8% decrement in FEVi following O3 exposure which was
17 significantly larger than mean FEVi decrements of 6.2, 7.9, and 8.3% in black females
18 and Caucasian males and Caucasian females, respectively. In the study by Seal et al.
19 (1993). there was potential that the increased FEVi decrements in blacks relative to
20 whites were due to increased O3 tissue doses since exercise rates were normalized to
21 BSA. Differences in O3 tissue doses between the races should not have occurred in the
22 Que et al. study, however, since exercise rates were normalized to lung volume (viz., 6-8
23 times FVC). Thus, the increased mean FEVi decrement in black males is not likely
24 attributable to systematically larger O3 tissue doses in blacks relative to whites.
25 Smokers are less responsive to O3 than nonsmokers. Spirometric and plethysmographic
26 pulmonary function decline, nonspecific airway hyperreactivity, and inflammatory
27 response of smokers to O3 were all weaker than data reported for nonsmokers. Although
28 all of these responses are intrinsically related, the functional association between them, as
29 in nonsmokers, has been weak. Similarly, the time course of development and recovery
30 of these effects as well their reproducibility was not different from nonsmokers. Chronic
31 airway inflammation with desensitization of bronchial nerve endings and an increased
32 production of mucus may plausibly explain the reduced responses to O3 in smokers
33 relative to nonsmokers (Frampton et al.. 1997b: Torres et al.. 1997).
34 The first line of defense against oxidative stress is antioxidants-rich ELF which
35 scavenges free radicals and limit lipid peroxidation. Exposure to O3 depletes the
36 antioxidant level in nasal ELF probably due to scrubbing of O3 (Mudway et al.. 1999a).
37 however, the concentration and the activity of antioxidant enzymes either in ELF or
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1 plasma do not appear to be related to O3 responsiveness (Samet et al., 2001; Avissar et
2 al.. 2000; Blomberg et al.. 1999). Carefully controlled studies of dietary antioxidant
3 supplementation have demonstrated some protective effects of a-tocopherol and
4 ascorbate on spirometric lung function from O3 but not on the intensity of subjective
5 symptoms and inflammatory response including cell recruitment, activation and a release
6 of mediators (Samet etal. 2001; Trengaetal.. 2001). Dietary antioxidants have also been
7 reported to attenuate O3-induced bronchial hyperresponsiveness in asthmatics (Trenga et
8 al.. 2001).
9 A number of studies(e.g., Romieu et al.. 2004a: David et al.. 2003; Corradi et al.. 2002;
10 Bergamaschi et al.. 2001) have reported that genetic polymorphisms of antioxidant
11 enzymes may modulate pulmonary function and inflammatory response to O3 challenge.
12 It appears that healthy carriers of NQO1 wild type in combination with GSTM1 null
13 genotype are more responsive to O3. Adults with GSTM1 null only genotype did not
14 show O3 hyperresponsiveness. In contrast, asthmatic children with GSTM1 null genotype
15 (Romieu et al.. 2004a) were reported to be more responsive to O3. However, in a
16 controlled exposure of mild-to-moderate asthmatics (n=23; 33 ± 11 years) to 300 ppb O3
17 for 2 hours with moderate exercise, Vagaggini et al. (2010) found that six of the subjects
18 had a NQO \wt and GSTM1 mill, but this genotype was not associated with the changes
19 in lung function or inflammatory responses to O3.
20 Kim et al. (2011) also recently reported that GSTM1 genotype was not predictive of
21 FEVi responses in young healthy adults (32 F, 27 M; 25.0 ± 0.5 year) that were roughly
22 half GSTM1-null and half GSTM1-sufficient. Sputum neutrophil levels, measured in a
23 subset of the subjects (13 F, 11 M), were also not significantly associated with GSTM1
24 genotype.
25 In a study of healthy volunteers with GSTM1 sufficient (n=19; 24 ± 3) and GSTM1 null
26 (n=16; 25 ± 5) genotypes exposed to 400 ppb O3 for 2 hours with exercise, Alexis et al.
27 (2009) found that inflammatory responses but not lung function responses to O3 were
28 dependent on genotype. At 4 hours post O3 exposure, both GSTM1 genotype groups had
29 significant increases in sputum neutrophils with a tendency for a greater increase in
30 GSTM1 sufficient than nulls. At 24 h postexposure, sputum neutrophils had returned to
31 baseline levels in the GSTM1 sufficient individuals. In the GSTM1 null subjects,
32 however, sputum neutrophil levels increased from 4 h to 24 h and were significantly
33 greater than both baseline levels and levels at 24 h in the GSTM1 sufficient individuals.
34 Since there was no FA control in the Alexis et al. (2009) study, effects of the exposure
3 5 other than O3 itself cannot be ruled out. In general, the findings between studies are
36 inconsistent. Additional studies that include control exposures are needed to clarify the
37 influence of genetic polymorphisms on O3 responsiveness.
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1 In a retrospective analysis of data from 541 healthy, nonsmoking, white males between
2 the ages of 18-35 years from 15 studies conducted at the U.S. EPA Human Studies
3 Facility in Chapel Hill, North Carolina, McDonnell et al. (2010) found that increased
4 body mass index (BMI) was associated with enhanced FEVi responses. The BMI effect
5 was of the same order of magnitude but in the opposite direction of the age effect where
6 by FEVi responses diminish with increasing age. In a similar retrospective analysis,
7 Bennett et al. (2007) found enhanced FEVi decrements following O3 exposure with
8 increasing BMI in a group of 75 healthy, nonsmoking, women (age 24 ± 4 years; BMI
9 range 15.7 to 33.4), but not 122 healthy, nonsmoking, men (age 25 ± 4 years; BMI range
10 19.1 to 32.9). In the women, greater O3-induced FEVi decrements were seen in
11 overweight (BMI >25) than in normal weight (BMI from 18.5 to 25), and in normal
12 weight than in underweight (BMI <18.5) (P trend < 0.022). Together, these results
13 indicate that higher BMI may be a risk factor for pulmonary effects associated with O3
14 exposure.
Repeated Ozone Exposure Effects
15 Based on studies reviewed in previous O3 AQCDs, several conclusions can be drawn
16 about repeated Ito 2 h O3 exposures. Repeated exposures to O3 causes enhanced (i.e.,
17 greater decrements) FVC and FEVi responses on the second day of exposure. The
18 enhanced response appears to depend to some extent on the magnitude of the initial
19 response (Horvath et al.. 1981). Small responses to the first O3 exposure are less likely to
20 result in an enhanced response on the second day of O3 exposure (Folinsbee et al.. 1994).
21 With continued daily exposures (i.e., beyond the second day) there is a substantial (or
22 even total) attenuation of pulmonary function responses, typically on the third to
23 fifth days of repeated O3 exposure. This attenuation of responses is lost in 1 week (Kulle
24 etal., 1982; Linn et al., 1982a) or perhaps 2 weeks (Horvath et al.. 1981) without O3
25 exposure. In temporal conjunction with pulmonary function changes, symptoms induced
26 by O3 (e.g., cough, pain on deep inspiration, and chest discomfort), are also increased on
27 the second exposure day and attenuated with repeated O3 exposure thereafter (Folinsbee
28 etal.. 1998; Foxcroft and Adams. 1986; Linnetal.. 1982a: Folinsbee et al.. 1980). In
29 longer-duration (4-6.6 hours), lower-concentration studies that do not cause an enhanced
30 second-day response, the attenuation of response to O3 appears to proceed more rapidly
31 (Folinsbee et al.. 1994).
32 Consistent with other investigators, Frank et al. (2001) found FVC and FEVi decrements
33 to be significantly attenuated following four consecutive days of exposure to O3 (250
34 ppb, 2 h). However, the effects of O3 on the small airways (assessed by a combined index
35 of isovolumetric FEF25_75, Vmax50 and Vmax75) showed a persistent functional reduction
36 from Day 2 through Day 4. Notably, in contrast to FVC and FEVi which exhibited a
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1 recovery of function between days, there was a persistent effect of O3 on small airways
2 function such that the baseline function on Day 2 through Day 4 was depressed relative to
3 Day 1. Frank et al. (2001) also found neutrophil (PMN) numbers in BAL remained
4 significantly higher following O3 (24 h after last O3 exposure) compared to FA.
5 Inflammatory markers from bronchioalveolar lavage fluid (BALF) following 4
6 consecutive days of both 2-h (Devlin et al.. 1997) and 4-h (Torres et al.. 2000; Christian et
7 al.. 1998) exposures have indicated ongoing cellular damage irrespective of the
8 attenuation of some cellular inflammatory responses of the airways, lung function and
9 symptoms response. These data suggest that the persistent small airways dysfunction
10 assessed by Frank et al. (2001) is likely induced by both neurogenic and inflammatory
11 mediators, since the density of bronchial C-fibers is much lower in the small than large
12 airways.
Summary of Controlled Human Exposure Studies on Lung Function
13 Responses in humans exposed to ambient O3 concentrations include: decreased
14 inspiratory capacity; mild bronchoconstriction; rapid, shallow breathing pattern during
15 exercise; and symptoms of cough and pain on deep inspiration (U.S. EPA. 2006b. 1996a).
16 Discussed in subsequent Sections 6.2.2.1 and 6.2.3.1, exposure to O3 also results in
17 airway hyperresponsiveness, pulmonary inflammation, immune system activation, and
18 epithelial injury (Que etal.; Mudway and Kelly. 2004a). Reflex inhibition of inspiration
19 results in a decrease in forced vital capacity and, in combination with mild
20 bronchoconstriction, contributes to a decrease in the FEVi. Healthy young adults exposed
21 to O3 concentrations > 60 ppb develop statistically significant reversible, transient
22 decrements in lung function if minute ventilation or duration of exposure is increased
23 sufficiently. With repeated O3 exposures over several days, FEVi and symptom responses
24 become attenuated in both healthy individuals and asthmatics, but this tolerance is lost
25 after about a week without exposure (Gong etal.. 1997a: Folinsbee et al.. 1994; Kulle et
26 al.. 1982). In contrast to the attention of FEVi responses, there appear to be persistent O3
27 effects on small airways function as well as ongoing cellular damage during repeated
28 exposures.
29 There is a large degree of intersubject variability in lung function decrements
30 (McDonnell. 1996). However, these lung function responses tend to be reproducible
31 within a given individual over a period of several months indicating differences in the
32 intrinsic responsiveness of individuals (Hazucha et al., 2003; McDonnell et al., 1985a). In
33 healthy young adults, O3-induced decrements in FEVi do not appear to depend on gender
34 (Hazucha et al.. 2003). body surface area or height (McDonnell et al.. 1997). lung size or
35 baseline FVC (Messineo and Adams. 1990). There is limited evidence that blacks may
36 experience greater O3-induced decrements in FEVi than age-matched whites (Que etal.;
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1 Seal et al., 1993). Healthy children experience similar spirometric responses but lesser
2 symptoms from O3 exposure relative to young adults (McDonnell et al.. 1985b). On
3 average, spirometric and symptom responses to O3 exposure appear to decline with
4 increasing age beyond about 18 years of age (McDonnell et al.. 1999; Seal et al.. 1996).
5 There is a tendency for slightly increased spirometric responses in mild asthmatics and
6 allergic rhinitics relative to healthy young adults (Torres etal.. 1996). Spirometric
7 responses in asthmatics appear to be affected by baseline lung function, i.e., responses
8 increase with disease severity (Horstman et al.. 1995).
9 Available information on recovery of lung function following O3 exposure
10 indicates that an initial phase of recovery in healthy individuals proceeds relatively
11 rapidly, with acute spirometric and symptom responses resolving within about 2 to 4 h
12 (Folinsbee and Hazucha. 1989). Small residual lung function effects are almost
13 completely resolved within 24 h. One day following O3 exposure, persisting effects on
14 the small airways assessed by decrements in FEF25_75 and altered ventilation distribution
15 have been reported (Frank etal.. 2001; Foster etal.. 1997).
6.2.1.2 Epidemiology
16 The O3-induced lung function decrements consistently demonstrated in controlled human
17 exposure studies (Section 6.2.1.1) provide biological plausibility for the epidemiologic
18 evidence presented in the 1996 and 2006 O3 AQCDs, in which short-term ambient O3
19 exposure was consistently associated with lung function decrements in diverse
20 populations (U.S. EPA. 2006b. 1996a). Coherence between the two disciplines was found
21 not only for effects observed in groups with higher expected personal O3 exposures and
22 higher exertion levels, including children attending summer camps and adults exercising
23 or working outdoors, but also for effects observed in children and individuals with pre-
24 existing respiratory disease such as asthma (U.S. EPA. 2006b. 1996a). Recent
25 epidemiologic studies focused more on children with asthma rather than on groups with
26 increased outdoor exposures or other healthy populations. Whereas a majority of recent
27 studies conducted in children with asthma indicated decreases in lung function in
28 association with increases in ambient O3 exposure, recent studies in adults with asthma
29 and individuals without asthma found both O3-associated decreases and increases in lung
30 function. Recent studies also provided additional data to assess whether particular lags of
31 O3 exposure were more strongly associated with decrements in lung function; whether O3
32 associations were confounded by copollutant exposures; and whether risk was affected by
33 factors such as corticosteroid (CS) use, genetic polymorphisms, elevated BMI, and diet.
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Table 6-1
Study
Korricketal.
(1998)
Thurston et al.
(1997)
Spektor et al.
(1988b)
Spektor et al.
(1988a)
Spektor and
Lippmann (1991)
Berryetal.
(1991)
Neas et al.
(1999)
Girardot et al.
(2006)
Selwyn etal.
(1985)
Thaller etal.
(2008)
Higginsetal.
Avol etal. (1990)
Burnett etal.
(1990)
Raizenne etal.
(1989)
Bra uer etal.
(1996)
Castillejosetal.
(1995)
Romieu et al.
(1998a)
Nickmilderet al.
(2007)
Brunekreef etal.
(1994)
Hoeketal.
(1993)
Braun-Fahrlander
etal. (1994)
Hoppe et al.
(1995): Hoppe et
al. (2003)
Chan etal.
(2005)
Mean and upper percentile concentrations of ozone in
epidemiologic studies examining lung function in populations with
increased outdoor exposures
Location
Mt. Washington,
NH
Connecticut
River Valley, CT
Tuxedo, NY
Fairview Lake,
NJ
Fairview Lake,
NJ
Hamilton, NJ
Philadelphia, PA
Great Smoky
Mountain NP, TN
Houston, TX
Galveston, TX
San Bernardino,
CA
Idyllwild, CA
Lake
Couchiching,
Ontario, CA
Lake Erie,
Ontario, CA
British Columbia,
Canada
Mexico City,
Mexico
Mexico City,
Mexico
Southern
Belgium
Netherlands
Wageningen,
Netherlands
Southern
Switzerland
Munich,
Germany
Taichung City,
Taiwan
Years/Season
1991,1992
Warm season
1991-1993
Warm season
1985
Warm season
1984
Warm season
1988
Warm season
July 1988
1993
Warm season
2002-2003
Warm season
1981
Warm season
2002-2004
Warm season
1987
Warm season
1988
Warm season
1983
Warm season
1986
Warm season
1993
Warm season
June 1990-
October1991
March-August
1996
2002
Warm season
1981
Warm season
1989
Warm season
1989
Warm season
1992
Warm season
2001
Cold season
Os Averaging Time
Hike-time avg
(2-1 2 h)
1-hmax
1 -h avg
1-havga
1-havga
1-h max
12-havg
(9:00 a.m.9:00 p.m.)
Hike-time avg
(2-9 h)
15-min max
1-hmax
1-havga
1-havga
1-havga
1-havga
1-h max
1-hmax
Work shift avg (6-1 2 h)
1-h max
8-h max
Exercise-time avg (10-
145min)
1-hmax
30-min avg
30-min max (1:00p.m.-
4:00 p.m.)
8-h avg
(9:00 a.m.-5:00 p.m.)
Mean/Median
Concentration
(ppb)
40
83.6
NR
53
69
NR
57.5 (Camp 1)
55.9 (Camp 2)
48. 1b
47
35 (median)
123
94
59
71
40
179
67.3
NR
42.8°
NR
NR
High 03 days: 65.9
Control 03 days: 27.2
35.6
Upper Percentile
Concentrations (ppb)
Max: 74
Max: 160
Max: 124
Max (1-h max): 113
Max (1-h max): 137
Max: 204
Max (Camp 1): 106
Max: 74.2b
Max: 135
Max: 118
Max: 245
Max: 161
Max: 95
Max (1-h max): 143
Max: 84
Max: 365
95th: 105.8
Max (across 6 camps): 24.5-112.7°
Max (across 6 camps): 18.9-81.1°
Max: 99.5°
Max: 122°
Max: 80°
Max (high 03 days): 86
Max: 65.1
Max = Maximum; NR = not reported
a1-h avg, preceding lung function measurement.
blndividual-level exposure estimates were derived based on time-activity diary data.
'Concentrations were converted from ug/m3 to ppb using the conversion factor of 0.51 assuming standard temperature (25°C) and pressure (1
atm).
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September 2011
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Populations with Increased Outdoor Exposures
1 Few epidemiologic studies characterizing acute O3-related respiratory morbidity have
2 accounted for time spent outdoors, which may be an important determinant of
3 interindividual variability in personal O3 exposure. Among epidemiologic studies, studies
4 of individuals engaged in outdoor recreation, exercise, or work are more comparable to
5 controlled exposure studies because of improved estimates O3 exposures, measurement of
6 lung function before and after discrete periods of outdoor activity, and examination of O3
7 effects during exertion when the dose of O3 reaching the lungs may be higher because of
8 higher ventilation and inhalation of larger volumes of air. Characteristics and ambient O3
9 concentration data from epidemiologic studies of populations with increased outdoor
10 exposures are presented in Table 6-1. Similar to findings from controlled human
11 exposure studies, the collective body of epidemiologic evidence clearly demonstrates
12 decrements in lung function in association with O3 exposures during periods of outdoor
13 activity or exercise of varying intensity and duration (15 minutes to 12 hours) (Figures 6-
14 3 to 6-5 and Tables 6-2 to 6-4).
Children Attending Summer Camps
15 Studies of children attending summer camps, most of which were discussed in the 1996
16 O3 AQCD, have provided important understanding of the impact of ambient O3 exposure
17 on respiratory effects in young, healthy children. These studies were noted for their on-
18 site measurement of ambient O3 and daily assessment of lung function by trained staff
19 over 1- to 2-week periods (Thurston et al.. 1997; Berry etal.. 1991; Spektor and
20 Lippmann. 1991; Avoletal. 1990; Burnett et al.. 1990; Higgins et al.. 1990; Raizenne et
21 al.. 1989: Spektor et al.. 1988a: Raizenne etal.. 1987V
22 In groups mostly comprising healthy children (ages 7-17 years), decrements in FEVi
23 were found to be associated consistently with ambient O3 exposures averaged over the
24 1-8 hours preceding lung function measurement (Figure 6-3 and Table 6-2). Kinney et al.
25 (1996) corroborated this association in a reanalysis combining 5367 lung function
26 measurements collected from 616 healthy children from six studies (Spektor and
27 Lippmann. 1991; Avoletal.. 1990; Burnett et al.. 1990; Higgins et al.. 1990; Spektor et
28 al.. 1988a; Raizenne et al.. 1987). Based on uniform statistical methods, a 40-ppb
29 increase in concurrent-hour O3 exposure was associated with a -20 ml (95% CI: -25, -14)
30 change in afternoon FEVi: (Kinney et al.. 1996). In these studies conducted in locations
1 To facilitate comparisons among epidemiologic studies, for all health endpoints in Chapter 6, effect estimates are presented in
terms of a standard increment in ambient O3 concentration, one for each of the three commonly examined O3 averaging times (1-h
max, 8-h max, and 24-h average). These standard increments are 40 ppb, 30 ppb, and 20 ppb for 1-h max, 8-h max, and 24-h avg
O3, respectively, and are based on annual mean to 95th percentile differences that are representative of measurements from
nationwide O3 monitors in U.S. Metropolitan Statistical Areas as described in detail in Section 7.1.3.2 of the 2006 O3AQCD (U.S.
EPA, 2006b).
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1
2
3
4
5
6
7
8
9
10
11
across the Northeast U.S. and Canada and California (Table 6-1) with varying pollutant
mix, a wide range in effect estimates was found. Study-specific effect estimates ranged
between a 0.76 and 48 ml decrease or a 0.3% to 2.2% decrease in study mean FE\V
Associations between ambient O3 exposure and peak expiratory flow (PEF) in camp
studies were more variable than were those with FEVi, as indicated by the wider range in
effect estimates and wider 95% CIs (Figure 6-3 and Table 6-2). Nonetheless, most effect
estimates indicated decreases in PEF in association with ambient O3 exposure. The
largest effect (mean 2.8% decline per 40-ppb increase in 1-h max O3) was estimated in a
group of campers with asthma (Thurston et al.. 1997). In this study, O3 also was
associated with increases in chest symptoms and bronchodilator use, suggesting that the
observed decreases in PEF may have been indicative of clinically significant effects.
Study
FEV, (mil
Spektoretal. (1988a)
Spektorand Lippmann
(1991)
Raizenne et al. (1987)
Burnett et al. (1990)
Higginsetal. (1990)
Avoletal. (1990)
Kinneyetal. (1996)
Berry etal. (1991)
Population
Camperswithout asthma <
Camperswithout asthma <
Camperswithout asthma -•-
Camperswithout asthma — • —
Camperswithout asthma — •—
Pooled estimate •••
Camperswithout asthma
-160 -120 -80 -40 0 40 80
Change in FEV1 (ml) per standardized increment in O3 (95% Cl)
PEF (ml/sec)
Spektoretal. (1988a)
Raizenne et al. (1987)
Burnett etal. (1990)
Higginsetal. (1990)
Avoletal. (1990)
Kinneyetal. (1996)
Berry etal. (1991)
Neasetal. (1999)
Thurston et al. (1997)
Camperswithout asthma
Camperswithout asthma
Camperswithout asthma
Camperswithout asthma
Camperswithout asthma
Pooled estimate
Camperswithout asthma
Camperswithout asthma
Campers with asthma
•*—•-
-160 -120 -80 -40 0 40 80
Change in PEF (ml/sec) per standardized increment in O3 (95% Cl)
Effect estimates are from single-pollutant models and are standardized to a 40-ppb increase for 1-h avg or 1-h max ozone
exposures and a 30-ppb increase for 12-h avg ozone exposures.
Figure 6-3 Changes in FEVi (ml) or PEF (ml/sec) in association with ambient
ozone exposure in studies of children attending summer camp.
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6-26
September 2011
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Table 6-2 Additional characteristics and quantitative data for studies
represented in Figure 6-3
Study
Location
Population
Standardized percent Standardized effect
change (95% Cl)a estimate (95% Cl)a
FEV,
(ml)
Spektor et al. (1988a) Fairview Lake, NJ
Campers without asthma
-0.93 (-1.5,-0.35)
-20.0 (-32.5, -7.5)
Fairview Lake, NJ
Campers without asthma
-2.2 (-3.1, -1.3)
-51. 6 (-72.8,-30.4)
Raizenne et al. (1989) ke Erie,
Campers without asthma
-0.48 (-0.80,-0.16)
-11.6 (-19.4,-3.8)
Burnett etal. (1990
Lake Couchiching,
Ontario, Canada
Campers without asthma
-0.32 (-1.8,1.2)
-7.6 (-42.1,26.9)
Higgins et al. (1990) San Bernardino, CA
Campers without asthma
-1.6 (-2.4,-0.87)
-33.6 (-49.3,-17.9)
Avol etal. (1991
Pine Springs, CA
Campers without asthma
-0.58 (-1.1,-0.12)
-12.8 (-23.0,-2.6)
Kinney et al. (1996) Pooled analysis
Campers without asthma
-0.90 (-1.2,-0.65)
-20.0 (-25.5,-14.5)
Berry etal. (1991)
Hamilton, NJ
Campers without asthma
Data not available
32.8 (6.9, 58.7)
PEF
(ml/sec)
Spektor et al. (1988a) Lake Fairview, NJ
Campers without asthma
-1.8 (-3.3,-0.40)
-80.0 (-142.7,-17.3)
Raizenne etal. (1989
Lake Erie, Ontario,
Canada
Campers without asthma
-0.07 (-0.56, 0.41)
-4.0 (-30.7, 22.7)
Burnett etal. (1990)
Lake Couchiching,
Ontario, Canada
Campers without asthma
-1.9 (-3.8, -0.05)
-106.4 (-209.9,-2.9)
Higgins et al. (1990) San Bernardino, CA
Campers without asthma
-0.87 (-2.1,-0.34)
-44.0 (-105,-17.2)
Avol etal. (1991)
Pine Springs, CA
Campers without asthma
1.9(0.71,3.1)
86.8(31.9,142)
Kinnevetal. (1996'
Pooled analysis
Campers without asthma
0.31 (-0.88,1.5)
6.8 (-19.1, 32.7)
Berry etal. (1991)
Hamilton, NJ
Campers without asthma
Data not available
-40.4 (-132.1,51.3)
Neasetal. (1999'
Philadelphia, PA
Campers without asthma
-0.58 (-1.5, 0.33)
-27.5 (-70.8, 15.8)
Thurston et al. (1997) CT River Valley, CT
Campers with asthma
-2.8 (-4.9, -0.59)
-146.7 (-261.7,-31.7)
aAII effect estimates are standardized to a 40-ppb increase in 1 -h avg or 1 -h max 03, except that from Neas et al. (1999). which is standardized
to a 30-ppb increase in 12-h avg (9:00 a.m.-9:00 p.m.) 03.
1 As has been observed in controlled human exposure studies, FEVi and PEF responses to
2 ambient O3 exposure varied among individual campers. Based on separate regression
3 analyses of data from individual subjects, O3 exposure was associated with a wide range
4 of changes in lung function across subjects (Berry etal.. 1991; Fliggins et al.. 1990;
5 Spektor et al.. 1988a). For example, in the study of children attending camp in Fairview
6 Lake, NJ, 36% of subjects had statistically significant O3-associated decreases in FEVi,
7 and the upper decile of response was a 6.3% decrease in FEVi per a 40-pbb increase in 1-
8 h avg O3 (Spektor et al.. 1988a).
9 In contrast with these previous studies, a recent cross-sectional study of children
10 attending six different summer camps in Belgium did not find an association between
11 ambient O3 exposure and lung function. The ambient O3 concentrations in this recent
12 study was in the range of those in previous studies (Table 6-1); however, this recent study
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1
2
3
4
differed from previous studies in that each subject was examined only on one day, and
investigators performed between-camp comparisons rather than within-subject
comparisons. Camps with higher daily 1-h max O3 concentrations did not consistently
have larger decreases in mean intraday FEVi or FEVi/FVC (Nickmilder et al.. 2007).
5
6
7
8
9
10
11
12
13
Populations Exercising Outdoors
Similar to camp studies, studies of individuals exercising outdoors were noted for the
serial examination of subjects over days with a wide range in ambient O3 concentrations
and onsite assessment of O3 exposures during discrete periods of outdoor exercise. These
studies collectively show that mean O3 exposures ranging from 40 to 66 ppb during
exercise of variable duration and intensity are associated with small (< 1 to 4% per
standardized increment in (V) decreases in lung function in adults (Figure 6-4 and Table
6-3). Similar observations were made in children exercising outdoors (Table 6-3). For
both adults and children, evidence was provided largely by older studies that were
reviewed in the 1996 and 2006 O3 AQCDs.
Study
Korricketal. (1998)
Girardotetal. (2006)
Hoppeetal. (2003)
Spektoretal. (1988b)
Brunekreefetal. (1994)
Population
Adults hiking
Adults hiking
Adults exercising
Adults exercising
Adults exercising
Exercise Duration
-4
-2
-1
Percentchangein FEV1 per standardized
incrementin O3 (95% Cl)
Figure 6-4 Percent change in FEVi in association with ambient ozone
exposures of adults exercising outdoors. Studies generally are
organized in order of decreasing exercise duration. Effect
estimates are from single-pollutant models and are standardized to
a 40-ppb increase for ozone exposures averaged over 15 minutes
to 1 hour and a 30-ppb increase for ozone exposures averaged over
3 to 8 hours.
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
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Table 6-3
Study
Additional characteristics and quantitative data for studies
represented in Figure 6-4 and results from studies in children
exercising outdoors
Location
Population
Exercise duration
Os Averaging
Time
Parameter
Standardized percent
change (95% Cl)a
Studies of adults
Korricketal. Q998)
Girardot et al.
(2006)
Hoppe et al. (2003)
Selwynetal.
M QR^b
> laoai
Spektor et al.
(1988a)b
Brunekreef etal.
(1994)
Mt. Washington,
NH
Great Smoky Mt,
TN
Munich, Germany
Houston, TX
Tuxedo, NY
Netherlands
Studies of children not included in
Braun-Fahrlander
etal. (1994)
Castillejosetal.
(1995)
Hoek etal. (1993)
Switzerland
Mexico City,
Mexico
Wageningen,
Netherlands
Adult day hikers
Adult day hikers
Adults exercising
Adults exercising
Adults exercising
Adults exercising
Figure 6-4
Children
exercising
Children
exercising
Children
exercising
2-1 2 h
2-9 h
2h
NR
15-55min
10min-1 h
10 min
15min (2 periods)
1 h
Hike duration
Hike duration
30-min max (1 :00
p.m. -4:00 p.m.)
15-min max
30-min avg
Exercise duration
30-min avg
1 -h avg
1 -h avg
FEV,
FEV,
FEV,
PEF
FEV,
FEV,
FEV,
PEF
FEV,
PEF
-1.5 (-2.8, -0.24)
0.72 (-0.46, 1 .90)
-1.3 (-2.6, 0.13)
-2.8 (-5.9, 0.44)
-16 ml (-31.1, -0.87)°
-1.31 (-2.0, -0.65)
-0.82 (-1.6, -0.02)
-3.8 (-6.9, -0.96)
-0.48 (-0.72, -0.24)
-2.2 (-4.9, 0.55)
NR= Not reported.
"Effect estimates are standardized to a 40-ppb increase for 03 exposures averaged over 15 min to 1 h and a 30-ppb increase for 03 exposures
averaged over 3 to 8 h.
bResults not included in the figure because data were not provided to calculate percent change in lung function.
The 95% Cl was constructed using a standard error that was estimated from the p-value
1 Two studies of adult day-hikers of similar design and ambient O3 concentrations
2 produced contrasting results (Girardot et al.. 2006; Korricketal.. 1998). These studies
3 mostly comprised white, healthy adults and examined changes in lung function associated
4 with O3 exposures during multihour (2-12 h) periods of outdoor exercise. Although
5 analyses of day-hikers were based on a one-time assessment of lung function, they
6 included much larger sample sizes compared with panel studies of individuals exercising
7 outdoors. Among 530 hikers on Mt. Washington, NH, Korrick et al. (1998) reported
8 posthike declines in FEVi and FVC of approximately 0.7-1.5% per a 30-ppb increase in
9 2- to 12-h avg O3. In contrast, among 354 hikers in Great Smoky Mountains National
10 Park, TN, Girardot et al. (2006) more recently found that O3 exposure was associated
11 with posthike increases in many of the same lung function indices. Several differences in
12 study characteristics were used by Girardot et al. (2006) to explain discrepant results,
13 including their use of a larger number of less-well trained technicians, shorter mean
14 duration of hike (5 hours versus 8 hours), and older mean age of their subjects.
15 As was observed in camp studies, the magnitudes of O3-associated decreases in lung
16 function varied among individual subjects. Korrick et al. (1998) found larger O3-
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1 associated decreases in FEVi among hikers who were male, had history of asthma or
2 wheeze, were never smokers, and hiked greater than 8 hours. Additionally, O3 was
3 associated with an increased odds of a greater than 10% decline in FEF25_75o/0 among
4 hikers (OR: 2.3 [95% CI: 1.2, 6.7] per 30-ppb increase in 2- to 12-h avg O3) (Korrick et
5 al.. 1998). Likewise, Hoppe et al. (2003) found that on days with 30-min max (1:00 p.m.-
6 4:00 p.m.) ambient O3 concentrations above 50 ppb, 14% of athletes had at least a 20%
7 decrease in lung function or 10% increase in airway resistance.
Outdoor Workers
8 The 2006 O3 AQCD indicated that ambient O3 exposure was associated consistently with
9 decrements in lung function among outdoor workers (U.S. EPA. 2006b). and recent
10 studies produced similar findings (Thaller et al., 2008; Chan and Wu. 2005) (Figure 6-5
11 and Table 6-4). Although most of these studies assessed O3 exposures using central site
12 measurements, they were noteworthy for the long periods of time spent outdoors (6-14
13 hours across studies). Further, associations between O3 exposure and lung function
14 decrements were found for time periods during which ambient O3 concentrations did not
15 exceed 80 ppb (Table 6-1) (Chan and Wu. 2005: Braueretal.. 1996: Hoppe etal.. 1995).
16 In particular, Many studies of outdoor workers found that in addition to same-day
17 exposures, O3 exposures lagged 1 or 2 days (Chan and Wu. 2005: Braueretal.. 1996) or
18 exposures averaged over 2 days (Romieu et al., 1998a) were associated with equal or
19 larger decrements in lung function (Figure 6-5 and Table 6-4).
20 Similar to other populations with increased outdoor exposure, the magnitudes of O3-
21 associated lung function decrements in outdoor workers were small. Per standardized
22 increment in O3 concentration1, decreases in lung function ranged between less than 1%
23 and 3.6%. The magnitude of decrease was not found to depend strongly on duration of
24 outdoor work or ambient O3 concentration. The largest decrease (6.4% per 40-ppb
25 increase in 1-h max O3) was observed among berry pickers in British Columbia who were
26 exposed to relatively low ambient O3 concentrations (work shift mean: 26.0 ppb [SD:
27 H-8]) but had longer periods of outdoor work (8-14 hours) (Braueretal.. 1996) (Figure
28 6-5 and Table 6-4). However, a much smaller O3-associated decrease in FEVi was found
29 among street workers in Mexico City who were exposed to higher O3 concentrations
30 (work shift mean: 67.3 ppb [SD: 24]) during a similar duration of outdoor work. Among
31 studies of outdoor workers, the smallest magnitude of decrease (0.4% decrease (95% CI:
32 -0.8, 0) in afternoon FEVi/FVC per 40-ppb increase in 1-h max O3) was observed among
33 lifeguards in Galveston, TX (Thaller et al.. 2008) whose outdoor work periods were
34 shorter than those of the berry pickers but who were exposed to a similar range of
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
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Study Population
Thaller et al. (2008) Lifeguards
Parameter O3 Lag Subgroup
FVC 0
FEV^FVC
Brauer et al. (1996) Berry pickers FEV, 0
1
Hoppe et al. (1995) Forestry workers FEV,
Romieu et al. (1998) Streetworkers FEV,
0 Placebo
Antioxidant supplement
0-1 avg Placebo
Antioxidant supplement
-7
-2 -1
Percentchange in lung function per standardized increment
in 03 (95% Cl)
Figure 6-5 Percent change in lung function in association with ambient ozone
exposures among outdoor workers. Studies generally are
organized in order of increasing mean ambient ozone
concentration. Effect estimates are from single-pollutant models
and are standardized to a 40-ppb increase for 30-min or 1-h max
ozone exposures.
Table 6-4 Additional characteristics and quantitative data for studies
represented in Figure 6-5
Study
Thaller et
al. (2008)
Braueretal.
(1996)
Hoppe et al.
(1995)
Romieu et
al. (1998a)
Chanetal.
(2005)"
Location
Galveston, TX
British
Columbia,
Canada
Munich,
Germany
Mexico City,
Mexico
Taichung City,
Taiwan
Population
Lifeguards
Berry pickers
Forestry
workers
Male street
workers
Mail carriers
Parameter
FVC
FEV,/FVC
FEV,
FEV,
FEV,
PEF
Duration of
outdoor work
6-8 h
8-14 h
NR
Mean (SD): 9 h
(1)
8h
O3 Averaging
Time
1-h max
1-h max
30-min max (1 :00
p.m. -4:00 p.m.)
1-h max
8-h avg (9:00 a.m.-
5:00 p.m.)
03
Lag
0
0
1
0
0
0-1
avg
0
1
Subgroup
Placebo
Antioxidant
Placebo
Antioxidant
Standardized percent
change (95% Cl)a
0.24 (-0.28, 0.72)
-0.40 (-0.80, 0)
-5.4 (-6.5, -4.3)
-6.4 (-8.0, -4.7)
-1.4 (-3.0, 0.16)
-2.1 (-3.3, -0.85)
-0.52 (-2.0, 0.97)
-3.4 (-6.0, -0.78)
-1.2 (-4 .2, 1.8)
-1 .0 (-1 .3, -0.66)
-1.1 (-1.5, -0.78)
NR= Not reported.
'Effect estimates are standardized to a 40-ppb increase for 30-min or 1-h max 03 and a 30-ppb increase for 8-h max 03.
aPEF results not included in figure.
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1
2
3
4
ambient O3 concentrations. Few studies provided information on ventilation rate or pulse
rate, thus it was difficult to ascertain whether differences in the magnitudes of O3-
associated decreases in lung function were related to differences in workers' levels of
exertion.
5
6
7
8
9
10
11
12
13
14
15
16
17
Associations at lower ozone concentrations
Studies of populations engaged in outdoor activity examined and found that associations
between O3 and lung function decrements persisted at lower O3 concentrations
(Table 6-5). Among adults exercising outdoors, Spektor et al. (1988b) found that
associations persisted in analyses restricted to 30-min max ambient O3 concentrations
less than 80 ppb, and for most lung function parameters, effect estimates were similar to
those obtained for the full range of O3 concentrations (Table 6-5). In a study of children
attending summer camp, similar effects were estimated for the full range of 1-h avg O3
concentrations and those less than 60 ppb (Spektor etal.. 1988a). Brunekreef et al. (1994)
found ambient O3 exposure (10-min to 1-h) during outdoor exercise to be associated with
decreases in lung function in analyses restricted to concentrations less than 61 (Table 6-5)
and 51 ppb. However, effect estimates were near zero with O3 concentrations less than
41 ppb (Brunekreef et al., 1994). In contrast, Brauer et al. (1996) found associations
persisted with 1-h max O3 concentrations less than 40 ppb.
Table 6-5
Study
Brunekreef etal. (1994)
Spektor et al. (1988b)
Spektor et al. (1988a)
Korrick etal. Q998)
Associations between ambient ozone exposure and lung function
decrements in different ranges of ambient ozone concentrations
Location Population
Netherlands Adults exercising
Tuxedo, NY Adults exercising
Fairview Lake, NJ Campers without
asthma
Mt. Washington, Adult day hikers
NH
Parameter
% change
FEV,
FEV, (ml)
% change
FEV,
% change
FEV,
03
Averaging
Time
10-mto1-h
LagO
30-min avg
LagO
1 -h avg
LagO
Hike duration
(2-12 h)
LagO
03
Concentration
Range
Full range
03 < 61 ppb
Full range
03 < 80 ppb
Full range
03 < 80 ppb
03 < 60 ppb
Full range
03 > 40 ppb
Standardized percent
change (95% Cl)a
-0.82 (-1 .6, -0.02)
-2.1 (-4.5, 0.32)
-54 (-84, -27)b
-52(-101,-3.4)b
-2.7 (-3.3, -2.0)
-1.4 (-2.5, -0.34)
-2.2 (-3.7, -0.80)
-1.5 (-2.8, -0.24)
-2.6 (-4.9, -0.32)
"Effect estimates are standardized to a 40-ppb increase for 03 exposures averaged over 10 min to 1 h and a 30-ppb increase for 03 exposures
averaged over 2 to12 h.
bData were not provided to calculate percent change.
18
19
20
Korrick et al. (1998) examined associations with hike-time average O3 exposures (2-12 h)
and found effect estimates that were more negative in analyses restricted to O3
concentrations greater than 40 ppb. Based on the results from a nonparametric model in
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1 Korrick et al. (1998). it appeared that the association between O3 exposure and lung
2 function decrements in this population was limited to 2- to 12-h avg O3 exposures above
3 40 ppb.
Children with Asthma
4 Associations between ambient O3 exposures and lung function decrements in children
5 with asthma have been examined in epidemiologic studies conducted across diverse
6 geographical locations and a range of ambient O3 concentrations (Table 6-6). Whereas
7 studies of populations with increased outdoor exposures monitored O3 exposures at the
8 site of subjects' outdoor activities and used trained staff to measure lung function, studies
9 of children with asthma relied more heavily on O3 measured at central monitoring sites
10 and lung function measured by subjects. However, studies of children with asthma have
11 provided more information on factors that may confer increased susceptibility to the
12 respiratory effects of O3 exposure, confounding by copollutant exposure or meteorology,
13 and the potential clinical significance of O3-associated changes in lung function with the
14 concurrent assessment of respiratory symptoms.
15 Collectively, the large body of evidence, which includes large U.S. multicity studies and
16 several smaller studies conducted in the U.S., Mexico City, and Europe, demonstrates
17 that increases in ambient O3 exposure (various averaging times and lags) are associated
18 with decrements in FEVi (Figure 6-6 and Table 6-7) and PEF (Figure 6-7 and Table 6-8)
19 in children with asthma. In addition to examining a single lung function measurement per
20 day, several studies examined associations of O3 exposure with measures of lung function
21 variability. Although different definitions of variability were used, studies consistently
22 found that O3-associated changes in lung function variability were indicative of poorer
23 lung function, whether characterized as a decrease from the individual's mean lung
24 function over the study period (Jalaludin et al.. 2000). a decrease in lung function over
25 the course of the day (Lewis et al.. 2005). or a decrease in the lowest daily measurement
26 (Just etal.. 2002).
27 Studies of children with asthma that were restricted to winter months provided little
28 evidence of an association between various single- and multi-day lags of ambient O3
29 exposure and changes in lung function; several studies reported O3-associated increases
30 in lung function (Dales et al.. 2009; Liu et al.. 2009a: Rabinovitch et al.. 2004). In colder
31 months, ambient O3 concentrations are low and in many locations, children remain
32 primarily indoors. Thus, it is less likely that effects will be demonstrated for O3. As noted
33 in previous AQCDs for lung function and other endpoints such as respiratory hospital
34 admissions, ED visits, and mortality, associations with O3 generally are greater in the
35 warm season.
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Table 6-6 Mean and upper percentile concentrations of ozone in
epidemiologic studies examining lung function in children with
asthma
Study
Mortimer etal. (2002)
Mortimer etal. (2000)
O'Connor etal. (2008)
Thurston et al. (1997)
Lewis etal. (2005)
Rabinovitchetal.
(2004)
Delfino et al. (2004)
Dales etal. (2009)
Liuetal. (2009a)
Romieu et al. (1996)
Romieu et al. (1997)
Romieu etal. (2002):
Romieu etal. (2004a):
Romieu et al. (2006)
Barraza-Villarreal et al.
(2008): Romieu etal.
(2009)
Hernandez-Cadena et
al. (2009)
Gielen etal. (1997)
Just etal. (2002)
Hoppe et al. (2003)
Wiwatanadate and
Trakultivakorn (2010)
Jalaludin et al. (2000)
Location
Bronx, East Harlem, NY; Baltimore,
MD; Washington, DC; Detroit, Ml,
Cleveland, OH; Chicago,!!.; St.
Louis, MO (NCICAS)
Boston, MA; Bronx, Manhattan NY;
Chicago, IL; Dallas, TX, Seattle,
WA; Tucson, AZ(ICAS)
Connecticut River Valley, CT
Detroit, Ml
Denver, CO
Alpine, CA
Windsor, ON, Canada
Northern Mexico City, Mexico
Southern Mexico City, Mexico
Mexico City, Mexico
Mexico City, Mexico
Mexico City, Mexico
Amsterdam, Netherlands
Paris, France
Munich, Germany
Chiang Mai, Thailand
Sydney, Australia
Years/Season
1993
Warm season
1998-2001
All-year
1991-1993
Warm season
2001-2002
All-year
1999-2002
Cold season
September-
October 1999
April-June 2000
2005
Cold season
April-July 1991
November 1991-
February 1992
April-July 1991
November 1991 -
February 1992
1998-2000
All-year
2003-2005
All-year
2005
Warm season
1995
Warm season
April-June 1996
1992-1995
Warm season
August 2005-June
2006
February-
December 1994
O3Averging
Time
8-h avg
(10:00a.m.-
6:00 p.m.)
24-h avg
1-h max
8-h max
1-h max
8-h max
24-h avg
1-h max
1-h max
1-h max
8-h max
1 -h max
8-h max
1 -h max
24-h avg
1-h max
8-h max
24-h avg
30-min max
(1:00 p.m. -4:00
p.m.)
24-h avg
15-h avg (6:00
a.m.-9:00p.m.)
Mean/Median
Concentration
(PPb)
48
NR
83.6a
Eastside: 40.4a
Westside:41.4a
28.2
62.9
14.1
27.2
190
196
69
102
31.6
86.5
26.3
74.5
34.2
30.0
High 03 days: 66.9
Control 03 days: 32.5
17.5
12
Upper Percentile
Concentrations
(PPb)
NR
NR
Max: 160
Overall max: 92.0a
Max 70.0
90th: 83.9, Max: 105.9
75th: 17.8
75th: 32.8
Max: 370
Max: 390
Max: 184
Max: 309
Max (8-h): 86.3
75th: 35.3; Max: 62.8
75th: 92.5; Max: 165.0
Max: 56.5
Max: 61 .7
Max: 91 (high 03 days)
39 (control 03 days)
90th: 26.82
Max: 34.65
Max: 43
NCICAS = National Cooperative Inner-City Asthma Study, NR = Not Reported, ICAS = Inner City Asthma Study, Max = Maximum.
'Measured at sites established by investigators.
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Study
Liu et al. (2009)
Lewis et al. (2005)
Hoppe et al. (2003)
O3Lag
0
2
Subgroup
CS user
With URI
1 Without asthma
With asthma
Barraza-Villarreal et al. (2008) 0-4 avg
Romieu et al. (2002)
Romieu et al. (2006)
Without asthma
With asthma
Placebo
Antioxidant
Placebo, moderate/severe asthma
Antioxidant, moderate/severe asthma
GSTP1 lie/lie Ile/Val
GSTP1 Val/Val
-10 -8 -6-4-20 2 4
Percent change in FEVi per standardized increment in O3 (95% Cl)
Figure 6-6 Percent change in FEVi in association with ambient ozone
exposures among children with asthma. Results generally are
presented in order of increasing mean ambient ozone
concentration. CS = Corticosteroid, URI = Upper respiratory
infection. Effect estimates are from single-pollutant models and are
standardized to a 40-ppb increase for 30-min or 1-h max ozone
exposures, a 30-ppb increase for 8-h max or 8-h avg ozone
exposures, and a 20-ppb increase for 24-h avg ozone exposures.
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Table 6-7 Additional characteristics and quantitative data for studies
represented in Figure 6-6
Study
Liu et al. (2009a)
Lewis etal. (2005)
Hoppe et al. (2003)
Barraza-Villarreal
etal. (2008)
Romieu et al.
(2002)
Romieu et al.
(2006)
Location/ Population
Windsor, ON, Canada
Children with asthma
Detroit, Ml
Children with asthma
Munich, Germany
Children
Mexico City, Mexico
Children
Mexico City, Mexico
Children with asthma
Mexico City, Mexico
Children with asthma
03
Averaging Os Lag
Time
24-h avg 0
8-h max 2
30-min max 1
(1:00p.m.-
4:00p.m.)
8-h max 0-4 avg
1-hmax 1
1-hmax 1
Parameter
FEV,
Lowest daily FEV,
Afternoon FEV,
Afternoon FVC
FEV,
FEV,
FEV,
Subgroup
CS user
With URI
Without asthma
With asthma
Without asthma
With asthma
Without asthma
With asthma
Placebo
Antioxidant
Placebo, moderate/severe
asthma
Antioxidant,
moderate/severe
asthma
GSTP1 lie/lie or Ile/Val
GSTP1 ValA/al
Standardized
percent change
(95% Clf
-0.89 (-3.5, 1 .8)
-8.0 (-13.5, -2.1)
-5.4 (-11. 3, 1.0)
0.93 (-0.80, 2.7)
-0.56 (-4.6, 3.7)
-0.09 (-1 .7, 1 .6)
-3.5 (-5.9, -1 .0)
-1 .5 (-4.7, 1 .7)b
-0.1 2 (-2.0, 1.8)"
-0.21 (-0.78, 0.36)b
0.05 (-0.59, 0.69)b
-1.1 (-2.0, -0.19)b
-0.04 (-0.92, 0.83)b
-0.51 (-1.1,0.05)
0.50 (-0.25, 1.3)
Studies not included in Figure 6-6b
Dales etal. (2009)
Rabinovitch etal.
(2004)
O'Connor etal.
(2008)
Windsor, ON, Canada
Children with asthma
Denver, CO
Children with asthma
7 U.S. communities
Children with asthma
1-hmax 0
1 -h max 0-2 avg
24-h avg 1 -5 avg
Evening %
predicted FEV,
Morning FEV, (ml)
Change in %
predicted FEV,
-0.47 (-1 .9, 0.95)
53 (-2.4, 108)
-0.41 (-1.0,0.21)
CS = corticosteroid, URI = Upper respiratory infection.
"Effect estimates are standardized to a 40-ppb increase for 30-min or
24-h avg 03.
°Results not presented in Figure 6-6 because a different form of FEV,
provided to calculate percent change in lung function.
1-h max 03, a 30-ppb increase for 8-h max 03, and a 20-ppb increase for
with a different scale was examined or because sufficient data were not
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Study Parameter
Gielenetal. (1997) PEF
Mortimeretal. (2002) PEF
Mortimeretal. (2000) PEF
Thurstonetal. (1997) PEF
Romieuetal. (2004) FEF26.76%
Romieu etal. (1996) Evening PEF
Romieu etal. (1997) Evening PEF
O3Lag Subgroup
2
1-5avg All subjects
Normal BW
LowBW
No asthma medication
CSuser
0
1 Placebo, GSTM1 null
Placebo, GSTM1 positive
Antioxidant, GSTM1 null
Antioxidant, GSTM1 positive
-10
-6
-4
Percent change in lung function parameter per standardized
increment in O3 (95% Cl)
Figure 6-7 Percent change in PEF or FEF25-75% in association with ambient
ozone exposures among children with asthma. Results generally
are presented in order of increasing mean ambient ozone
concentration. BW = birth weight, CS = Corticosteroid. Effect
estimates are from single pollutant models and are standardized to
a 40-ppb increase for 1-h max ozone exposures and a 30-ppb
increase for 8-h max or 8-h avg ozone exposures.
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Table 6-8
Study
Gielenetal. (1997)
Mortimer etal. (2002)
Mortimer etal. (2000)
Thurston et al. (1997)
Romieu et al. (2004a)
Romieu et al. (1996)
Romieu et al. (1997)
Studies not included
Jalaludin et al. (2000)
Wiwatanadate and
Trakultivakorn (2010)
O'Connor etal.
(2008)
Additional characteristics and quantitative data for studies
represented in Figure 6-7
Location/ Population
Amsterdam, Netherlands
Children w/asthma
8 U.S. communities
Children w/asthma
8 U.S. communities
Children w/asthma
CT River Valley, CT
Children w/asthma
Mexico City, Mexico
Children w/asthma
Northern Mexico City, Mexico
Children w/asthma
Southern Mexico City, Mexico
Children w/asthma
in Figure 6-7b
Sydney, Australia
Children w/asthma
Chiang Mai, Thailand
Children w/asthma
7 U.S. communities
Children w/asthma
03
Averaging
Time
8-h max
8-h avg
(10:00a.m.-
6:00p.m.)
8-h avg
(10:00a.m.-
6:00p.m.)
1 -h avg
1-h max
1-hmax
1-h max
24-h avg
24-h avg
24-h avg
03
Lag
2
1-5
avg
1-5
avg
0
1
0
2
0
2
0
0
5
1-5
avg
Parameter Subgroup
PEF
PEF All subjects
PEF Normal BW
LowBW
No medication
CS user
I ntraday change
PEF
FEF2™ Placebo, GSTM1 null
Placebo, GSTM1 positive
Antioxidant, GSTM1 null
Antioxidant, GSTM1 positive
Evening PEF
Evening PEF
% variability Wheeze, no asthma
PEF Asthma, no AHR
Asthma, with AHR
Daily avg PEF
(L/min)
Change in %
predicted PEF
Standardized
percent change
(95% Cl)a
-1.3 (-2.6, -0.10)
-1.2 (-2.1, -0.26)
-0.60 (-1 .6, 0.39)
-3.6 (-5.2, -2.0)
-1.1 (-3.0,0.84)
-1.2 (-2.5, 0.11)
-2.8 (-4.9, -0.59)
-2.3 (-4.2, -0.44)
-0.48 (-1.7, 0.74)
-0.1 6 (-1.8, 1.6)
0.24 (-1.3, 1.8)
-0.1 7 (-0.79, 0.46)
-0.55 (-1.3, 0.1 9)
-0.52 (-1.0, -0.007)
-0.06 (-0.70, 0.58)
3.8 (0.25, 7.38)°
-0.71 (-2.6, 1 .2)°
-5.2 (-8.3, -2.2)°
1 .0 (-1 .6, 3.6)
-2.6 (-5.2, 0)
-0.22 (-0.86, 0.43)
BW = birth weight, CS = corticosteroid, AHR = Airway hyperresponsiveness.
"Effect estimates are standardized to a 40-ppb increase for 1-h max 03, a 30-ppb increase for 8-h max or 8-h avg 03, and a 20-ppb increase for
24-h avg 03.
bResults are not presented in Figure 6-7 because a different form of PEF with a different scale was examined or because sufficient data were not
provided to calculate percent change in lung function.
0 Outcome defined as the percent deviation from individual mean PEF during the study period. Group-stratified effect estimates were provided
only for models that included PM10 and N02.
1 The most geographically representative data were provided by the large, multi-U.S. city
2 National Cooperative Inner City Asthma Study (NCICAS) (Mortimer et al., 2002;
3 Mortimer et al.. 2000) and Inner-City Asthma Study (ICAS) (O'Connor et al.. 2008V
4 Although the two studies differed in the cities, seasons, racial distribution of subjects, and
5 lung function indices examined, results were fairly similar. In ICAS, which included
6 children with asthma and atopy (i.e., allergic sensitization) and year-round examinations
7 of lung function, a 20-ppb increase in the lag 1-5 average of 24-h avg O3 was associated
8 with a 0.41-point decrease in percent predicted FEVj (95% CI: -1.0, 0.21) and a 0.22-
9 point decrease in percent predicted PEF (95% CI: -0.86, 0.43).
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1 Lag 1-5 avg O3 (8-h avg, 10:00 a.m.-6:00 p.m.) also was associated with declines in PEF
2 in NCICAS, which included different U.S. cities, summer-only measurements, larger
3 proportions of Black and Hispanic children, and fewer subjects with atopy (79%)
4 (Mortimer et al.. 2002). NCICAS additionally identified groups potentially at increased
5 risk of O3-associated decrements in PEF. Larger effects were estimated in males, children
6 of Hispanic ethnicity, children living in crowded housing, and as indicated in Figure 6-7
7 and Table 6-8, children with low birth weight (Mortimer et al., 2000). Somewhat
8 paradoxically, O3 was associated with a larger decrease in PEF among subjects taking
9 cromolyn, medication typically used to treat asthma due to allergy, but a smaller decrease
10 among subjects with positive atopy (as determined by skin prick test). Similar to
11 observations from studies of populations with increased outdoor exposures, Mortimer et
12 al. (2002) found that associations persisted at lower ambient O3 concentrations. At
13 concentrations below 80 ppb, a 30-ppb increase in lag 1-5 of 8-h avg O3 was associated
14 with a 1.4% decrease (95% CI: -2.6, -0.21) in PEF, which was similar to the effect
15 estimated for the full range of O3 concentrations (Figure 6-7 and Table 6-8). In a study of
16 children with asthma in the Netherlands, Gielen et al. (1997) estimated similar effects for
17 the full range of 8-h max O3 concentrations and concentrations below 51 ppb.
18 The results from studies of children with asthma indicated that factors in addition to
19 asthma influenced associations between ambient O3 exposure and changes in lung
20 function. In comparisons between children with and without asthma, Hoppe et al. (2003)
21 and Jalaludin et al. (2000) generally found larger O3-associated lung function decrements
22 in children with asthma; whereas Raizenne et al. (1987) did not consistently demonstrate
23 differences between campers with and without asthma. In their study of children in
24 Mexico City, Barraza-Villarreal et al. (2008) estimated larger O3-associated decreases in
25 children without asthma; however, 72% of these children had atopy. These findings
26 indicated that in addition to asthma, atopy, a condition also characterized by airway
27 inflammation and similar respiratory symptoms, may increase the risk for O3-associated
28 respiratory effects.
29 As indicated in Figures 6-6 and 6-7 and Tables 6-7 and 6-8, in most studies of children
30 with asthma, standardized increments in ambient O3 exposure1 were associated with
31 decreases in lung function that ranged from less than 1% to 2%. Larger magnitudes of
32 decreases (3-8% per standardized increments in O3) were found in children with asthma
33 who also were using CS, had a concurrent upper respiratory infection (URI), were
34 GSTM1 null, had low birth weight, or had increased outdoor exposure (Romieu et al.,
35 2006; Lewis etal.. 2005; Romieu et al.. 2004a: Jalaludin et al.. 2000) than among
36 children with asthma overall (Barraza-Villarreal et al.. 2008; Lewis etal.. 2005; Delfino
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
Draft - Do Not Cite or Quote 6-39 September 2011
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1 et al.. 2004; Romieu et al.. 2002). For example, Jalaludin et al. (2000) estimated a -5.2%
2 deviation from mean FEVi per a 20-ppb increase in 24-h avg O3 among children with
3 asthma and airway hyperresponsiveness and a much smaller -0.71% deviation among
4 children with asthma without airway hyperresponsiveness. In a group of 86 children with
5 asthma in Detroit, MI, Lewis et al. (2005) also reported that associations between O3
6 exposure and lung function decrements were confined largely to children with asthma
7 who used CS or had a concurrent URL These two groups were observed to have the
8 largest O3-associated decrements in lung function among all studies of children with
9 asthma. A 30-ppb increase in 8-h max ambient O3 exposure was associated with a 8.0%
10 decrease in the mean of lowest daily FEVi among CS users and a 5.4% decrease among
11 subjects reporting concurrent URI (Lewis et al.. 2005) (Figure 6-6 and Table 6-7).
12 Heterogeneity in lung function responses to ambient O3 exposure also has been
13 demonstrated as inter-individual variability in the magnitude of O3-associated changes in
14 lung function. Mortimer et al. (2002) found that for a 30-ppb increase in lag 1-5 avg of 8-
15 h avg O3, there was a 30% (95% CI: 4, 61) increase in the incidence of a greater than
16 10% decline in PEF. Likewise, Hoppe et al. (2003) found that while the percentages of
17 change in individual lung function parameters were variable and small, 47% of children
18 with asthma in their study experienced greater than 10% decline in FEVi, FVC, or PEF
19 or 20% increase in airway resistance on days with 30-min (1:00 p.m.-4:00 p.m.) max
20 ambient O3 concentrations greater than 50 ppb relative to days with less than 40 ppb O3.
21 In addition to finding groups of children with asthma with increased sensitivity to O3
22 exposure, epidemiologic studies have indicated that the decreases in lung function
23 observed in association with increases in ambient O3 exposure may be clinically
24 significant by finding that the same or similar lag of O3 exposure was associated with
25 decrements in lung function and increases in concurrently assessed respiratory symptoms
26 (Just et al.. 2002: Mortimer etal.. 2002: Gielenetal.. 1997: Romieu etal.. 1997:
27 Thurston et al.. 1997; Romieu et al.. 1996) (see Figure 6-12 and Table 6-19 for symptom
28 results).
Effect modification by corticosteroid use
29 In controlled human exposure studies, CS treatment of subjects with asthma generally has
30 not prevented O3-induced FEVi decrements (Section 6.2.1.1). In epidemiologic studies
31 reviewed in the 2006 O3 AQCD, evidence was equivocal, as use of inhaled CS showed
32 both protective (Delfino et al.. 2002; Mortimer et al.. 2000) and exacerbating (Gent et al..
33 2003) effects on respiratory endpoints. Among recent studies, evidence for effect
34 modification of lung function responses by CS use also was mixed. In Lewis et al.
35 (2005). analyses of interactions between O3 exposure and CS use indicated stronger
36 associations among CS users than among CS nonusers (quantitative results not reported
Draft - Do Not Cite or Quote 6-40 September 2011
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1 for CS nonusers). Among the 11 (12.8%) CS users, a 30-ppb increase in lag 2 of 8-h max
2 O3 was associated with an 8.0% decrease (95% CI: -13.5, -2.1) in lowest daily FEVi and
3 a 6.7% increase (95% CI: 0.60, 13.2) in diurnal FEVi variability (indicating a decrease
4 from morning to evening). Other lags (1 or 3-5 avg) or averaging times (24-h avg) of
5 exposure were estimated to have less impact. In contrast to Lewis et al. (2005),
6 Hernandez-Cadena et al. (2009) observed greater O3-related decrements in FEVi among
7 the 60 CS nonusers than among the 25 CS users. In two winter-only studies,
8 consideration of CS use did not largely influence associations between ambient O3 and
9 lung function parameters (Liu et al.. 2009a; Rabinovitch et al.. 2004).
10 Although studies varied in populations and season examined, the inconsistency in effect
11 modification by CS use may be explained, at least in part, by differences in the severity
12 of asthma among CS users and the definition of CS use. Hernandez-Cadena et al. (2009)
13 did not define CS use; however, the group of CS nonusers included both children with
14 intermittent and persistent asthma. In Lewis et al. (2005). most children with moderate to
15 severe asthma (91%) were included in the group of CS users (use for at least 50% of
16 study days); however, these subjects had a higher percent predicted FEVi • Liu et al.
17 (2009a) did not provide information on asthma severity; however, they defined CS use
18 more stringently as daily use. Differences in asthma severity and definition of CS use
19 may explain why both CS use and nonuse could serve as indicators of severe or
20 uncontrolled asthma across studies. Additionally, investigators did not assess adherence
21 to reported CS regimen, and misclassification of CS use may bias findings.
Effect modification by antioxidant capacity
22 Ozone is a powerful oxidant whose secondary oxidation products are recognized to
23 initiate the key modes of action, including the activation of neural reflexes that mediate
24 decreases in lung function (Section 5.3.2). Additionally, O3 exposure of humans and
25 animals induces changes in the levels of antioxidants in the ELF (Section 5.3.3). These
26 observations support the biological plausibility for diminished antioxidant capacity
27 increasing the risk of O3-associated respiratory effects and augmented antioxidant
28 capacity decreasing risk. Controlled human exposure studies have demonstrated
29 protective effects of a-tocopherol (vitamin E) and ascorbate (vitamin C) supplementation
30 on O3-induced lung function decrements (Section 6.2.1.1), and epidemiologic studies of
31 children with asthma conducted in Mexico City have had similar findings. Particularly
32 among children with moderate to severe asthma, ambient O3 exposure was associated
33 with a smaller decrease in FEVi in the group supplemented with vitamin C and E as
34 compared with the placebo group (Romieu et al.. 2002) (Figure 6-6 and Table 6-7).
35 Similarly, Romieu et al. (2009) observed protective effect for diets high in vitamins C
36 and E as well as omega-3 fatty acids. Subjects were assigned to a fruits and vegetables
Draft - Do Not Cite or Quote 6-41 September 2011
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1 index (FVI) that characterized consumption of vitamins C and E and a Mediterranean diet
2 index (MDI) that additionally represented the intake of omega-3 fatty acids, which have
3 anti-inflammatory effects. At lag 0-4 avg of 8-h max O3 concentrations > 38 ppb, a 1-unit
4 increase in FVI was associated with a 137 ml (95% CI: 8, 266) increase in FEVi. This
5 protective effect of FVI was diminished at O3 concentrations < 25 ppb (65 ml [95% CI: -
6 70, 200] increase in FEVi per 1-unit increase in FVI). Similar results were obtained for
7 MDI.
8 Antioxidant capacity also can be characterized by variants in genes encoding xenobiotic
9 metabolizing enzymes with different enzymatic activities. Ambient O3 exposure has been
10 associated with greater decreases in lung function among children with asthma with the
11 GSTM1 null genotype, which is associated with lack of oxidant metaboilizing activity
12 (Romieu et al.. 2004a). The difference in response between GSTM1 null and positive
13 subjects was minimal in children supplemented with antioxidant vitamins (Figure 6-7 and
14 Table 6-8). Although these findings are biologically plausible given the well-
15 characterized evidence for O3 effects mediated by secondary oxidation products, it is
16 important to note that a larger body of controlled human exposure studies has not
17 consistently found larger O3-induced lung function decrements in GSTM1 null subjects
18 (Section 6.2.1.1). Effect modification by the GSTP1 variant is less clear. Romieu et al.
19 (2006) observed larger O3-associated decreases in FEVi in children with asthma with the
20 GSTP1 lie/lie or Ile/Val variant, both of which are associated with normal oxidative
21 metabolism activity (Figure 6-6 and Table 6-7). Also unexpectedly, O3 exposure was
22 associated with an increase in FEVi among children with the GSTP1 Val/Val variant,
23 which is associated with reduced oxidative metabolism. Rather than reflecting effect
24 modification by the GSTP1 variant, these results may reflect effect modification by
25 asthma severity, as 77% of subjects with the GSTP1 lie/lie genotype had moderate to
26 severe asthma. Supporting evidence is provided by an earlier analysis of the same cohort,
27 in which the effect of antioxidant supplementation was demonstrated more strongly in the
28 smaller group of children with moderate to severe asthma than among all subjects with
29 asthma (Romieu et al., 2002).
Adults with Respiratory Disease
30 Relative to studies in children with asthma, studies of adults with asthma or COPD have
31 been limited in number. Characteristics and ambient O3 concentration data from these
32 studies are presented in Table 6-9. Studies that included both children and adults with
33 asthma did not consistently demonstrate associations between ambient O3 exposure and
34 decrements in lung function (Ross et al., 2002; Delfino et al., 1997). Ross et al. (2002)
35 found that a 20-ppb increase in lag 0 of 24-h avg O3 was associated with a 2.6 L/min
36 decrease (95% CI: -4.3, -0.90) in evening PEF among subjects ages 5-49 years. This
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1 decrement may have been indicative of a clinically significant effect, as lag 0 O3
2 exposure also was associated with an increase in symptom score. In another panel study,
3 neither ambient nor personal O3 12-h avg exposure was reported to be associated with a
4 decrease in lung function among subjects ages 9-46 years (Delfino et al.. 1997).
5 Comparisons of adults with and without asthma did not conclusively demonstrate that
6 adults with asthma were at increased risk of O3-associated respiratory effects. In the
7 recent panel study of 16- to 27-year-old lifeguards in Galveston, TX, a larger O3-
8 associated decrement in FEVi/FVC was found among the 16 lifeguards with asthma (-
9 1.6% [95% CI: -2.8, -0.4] per 40 ppb increase in 1-h max O3) than among the 126
10 lifeguards without asthma (-0.40% [95% CI: -0.80, 0] per 40 ppb increase in 1-h max O3)
11 (Brooks. 2010). In the studies of day-hikers, Korrick et al. (1998) found that the O3-
12 associated lung function decrements observed among all hikers were driven by
13 associations observed in hikers with history of asthma or wheeze (-4.4% [95% CI: -7.5, -
14 1.2] in FEVi per 30-ppb increase in 2-9 hr avg O3). In contrast, Girardot et al. (2006) did
15 not find ambient O3 exposure to be consistently associated with decrements in lung
16 function in subjects with or without respiratory disease history. In another cross-sectional
17 study of 38 adults with asthma and 13 adults without asthma, atopy was observed to be a
18 stronger susceptibility factor than was asthma (Khatri etal. 2009). Investigators reported
19 a larger decrease in percent predicted FEVi/FVC per 30-ppb increase in lag 2 of 8-h max
20 O3 among the 38 subjects with atopy (with or without asthma) (-12 points [95% CI: -21, -
21 3]) than among subjects with asthma (-4.7 points [95% CI: -11, 2.3]). Additionally,
22 among adults with asthma, O3 was associated with an increase in FEVi • Based on
23 correlations observed between decreases in lung function and decreases in quality of life
24 scores, investigators inferred the O3-associated decreases in lung function to be clinically
25 significant. They suggested that atopy may influence responses to ambient O3 exposure
26 because during the summer, high ambient O3 concentrations may increase the
27 allergenicity of pollens.
28 O3 was not found to have a strong effect on the lung function of adults with asthma in
29 panel studies conducted in Europe and Asia during low ambient O3 periods
30 (Wiwatanadate and Liwsrisakun. 2011; Lagorio et al., 2006; Park et al., 2005a). including
31 one study conducted in Korea during a period of dust storms (Park etal.. 2005a). In these
32 studies that examined multiple lags of O3 exposure, O3 generally was associated with
33 increases in lung function.
34 Controlled human exposure studies demonstrate robust O3-induced spirometric responses
35 in children and young adults but diminished, statistically nonsignificant responses in
36 older adults, both healthy and with COPD (Section 6.2.1.1). Similarly, in a recent
37 epidemiologic study that followed 94 adults with COPD (ages 40-83 years) daily over a
Draft - Do Not Cite or Quote 6-43 September 2011
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
2-year period, an increase in ambient O3 exposure was not associated consistently with
decreases PEF, FEVi, and FVC (Peacock et al.. 2011). For example, in an analysis
restricted to the summer of 1996, a 30-ppb increase in 8-h max O3 was associated with a
1.7 L/min decrease (95% CI: -3.1, -0.39) in PEF. However, during the summer of 1997,
O3 was found to have little effect on PEF (-0.21 L/min [95% CI: -2.4, 2.0] per 30-ppb
increase in 8-h max O3). Further, in this study, an increase in ambient O3 exposure was
associated with a lower odds of a large PEF decrement (OR for a greater than 20% drop
from an individual's median value: 0.89 [95% CI: 0.72, 1.10] per 30-ppb increase in lag 1
of 8-h max O3) and was not consistently associated with increases in respiratory
symptoms (Peacock et al.. 2011). Ozone exposure also was not consistently associated
with decreases in lung function in a smaller panel study of 11 adults with COPD (mean
age 67 years) (Lagorio et al.. 2006). Together, these finding do not provide strong
evidence that increases in O3 exposure are associated with lung function decrements in
adults with COPD.
Table 6-9
Study
Korricket al.
(1998)
Khatrietal. (2009)
Ross et al. (2002)
Thaller etal.
(2008)
Delfino et al.
(1997)
Lagorio etal.
(2006)
Peacock etal.
(2011)
Wiwatanadate et
al. (2011)
Park etal. (2005a)
Mean and upper percentile concentrations of ozone in
epidemiologic studies examining lung function in adults with
respiratory disease
Location
Mt.
Washington,
NH
Atlanta, GA
East Moline, IL
Galveston, TX
Alpine, CA
Rome, Italy
London,
England
Chiang Mai,
Thailand
Incheon, Korea
Years/Season
1991,1992
Warm season
2003, 2005, 2006
Warm season
April-October 1994
2002-2004
Warm season
1994
Warm season
1999
Spring and winter
1995-1997
All-year
August 2005 -
June 2006
March-June 2002
O3 Averaging
Time
Hike-time avg
(2-1 2 h)
8-h max
8-h avg
1-hmax
12-havg
personal
(8:00 a.m.-8:00
p.m.)
24-h avg
8-h max
24-h avg
24-h avg
Mean/Median Upper Percentile
Concentration (ppb) Concentrations (ppb)
40
59 (median)8
41.5
35 (median)
18
Spring: 36.2"
Winter: 8.0b
15.5
17.5
Dust event days: 23.6
Control days: 25.1
Max: 74
75tn: 73a
Max: 78.3
Max: 118
90th: 38
Max: 80
Overall max: 48.6"
Autumn/Winter Max: 32
Spring/Summer Max: 74
90th: 26.82
Max: 34.65
NR
NR = Not reported, Max = Maximum.
'Individual-level exposure estimates were derived based on time spent in the vicinity of various 03 monitors.
bConcentrations converted from ug/m3 to ppb using the conversion factor of 0.51 assuming standard temperature (25°C) and pressure (1 atm).
15
16
Populations Not Restricted to Individuals with Asthma
Several studies have examined associations between ambient O3 exposure and lung
function in children; however, a limited number of studies have examined other
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6-44
September 2011
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1
2
3
populations not restricted to individuals with asthma or other healthy populations.
Characteristics and ambient O3 concentration data from studies not restricted to
individuals with asthma are presented in Table 6-10.
Table 6-10
Study
Alexeefetal. (2007)
Alexeefetal. (2008)
Naeheretal. (1999)
Avoletal. (1998a)
Linn et al. (1996)
Gold et al. (1999)
Scarlett etal. (1996)
Ward etal. (2002)
Ulmer etal. (1997)
Hoppe et al. (2003)
Neuberger etal. (2004)
Steinvil etal. (2009)
Chen etal. (1999)
Son et al. (2010)
Mean and upper percentile concentrations of ozone in
epidemiologic studies examining lung function in populations not
restricted to individuals with asthma
Location
Greater Boston, MA
Vinton, VA
6 southern CA
communities
Rubidoux, Upland,
Torre nee, CA
Mexico City, Mexico
Surrey, England
Birmingham and
Sandwell, England
Freudenstadt and
Villingen, Germany
Munich, Germany
Vienna, Austria
Tel Aviv, Israel
3 Taiwan
communities
Ulsan, Korea
Years/Season
1995-2005
All-year
1995-1996
Warm season
1994
Spring and summer
1992-1993,1993-
1994
Fall and spring
1991
Winter, spring, fall
1994
Warm season
1997
Winter and summer
1994
March-October
1992-1995
Warm season
June-October 1999,
January-April 2000
2002-2007
All-year
1995-1996
May-January
2003-2007
All-year
Metric
24-h avg
8-h max
24-h avg personal
24-h avg
24-h avg
8-h max
24-h avg
30-min max
30-min max (1 :00 p.m.-
4:00 p.m.)
NR
8-h avg
(10:00 a.m. -6:00 p.m.)
1-hmax
8-h max
Mean/Median
Concentration (ppb)
24.4a
34.87
NR
23
52.0
50. 7b
Winter median: 13.0
Summer median: 22.0
Freudenstadt median: 50.6
Villingen median: 32.1
High 03 days: 70.4
Control 03 days: 29.8
NR
41.1
NR
35.86
Upper Percentile
Concentrations (ppb)
NR
Max: 56.63
NR
Max: 53
Max: 103
Max: 128b
Winter Max: 33
Summer Max: 41
Freudenstadt 95th:
Villingen 95th: 70.1
89.7
Max (high 03 days): 99
Max (control 03 days): 39
NR
75th: 48.7
Max: 72.8
Max: 110.3b
Max: 59.53
NR = Not Reported, Max = Maximum.
'Measured at central monitoring sites established by investigators. Concentations were averaged across all monitors.
bMeasured at subjects' schools where lung function measurements were performed.
Children
4 The 2006 O3 AQCD identified children as a potentially at-risk population based on
5 consistent evidence of association between ambient O3 exposure and decrements in
6 and PEF (U.S. EPA. 2006b) (Figure 6-8 and Table 6-11). No new studies in children
7 without asthma are available to compare with previous findings. Hoppe et al. (2003) O3
8 exposure to be associated with decreses in healthy children in Munich, Germany (Figure
9 6-8 and Table 6-11). In another panel study of healthy children in Vienna, Austria, O3
10 was not associated with decrements in total lung capacity (Neuberger et al.. 2004). Most
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1
2
3
4
5
6
7
8
9
10
11
of the studies in children did not exclusively examine healthy children. However, several
studies of children that included small proportions (4- 10%) of children with history of
respiratory disease or symptoms and found associations between O3 exposure and
decrements in lung function (Chen et al.. 1999; Ulmeretal.. 1997; Scarlett et al.. 1996).
Based on interactions between O3 exposure and asthma/wheeze history, Avol et al.
(1998a) and Ward et al. (2002) did not find lung function responses to ambient O3
exposure to differ between children with history of asthma or wheeze and healthy
children. Combined, these lines of evidence indicate that the associations observed
between ambient O3 exposure and decreases in lung function in children are not driven by
effects in children with asthma or respiratory symptoms, and that healthy children also
may represent a population at increased risk of O3-associated respiratory effects.
Parameter
Study
Linnetal. (1996)
Hoppeetal. (2003)
Scarlettetal. (1996)
Chenetal.(1999) FEV1
O3 Lag
Intraday change F£V! 0
Intraday change FVC 0
FEV.,
FVC
0
1
Avoletal. (1998)a Intraday change FEV., O(personal)
Intraday change FVC
-10
-6
-2
0
Percentchangein lung function parameter per standardized
increment in O3 (95% Cl)
Results generally are presented in order of increasing mean ambient ozone concentration.
aThe 95% Cl was constructed using a standard error that was estimated from the p-value. Effect estimates are from single-
pollutant models and are standardized to a 40-, 30-, and 20-ppb increase for a 1-h (or 30-min) max, 8-h max, and 24-h avg ozone
exposures, respectively.
Figure 6-8 Percent change in lung function in association with ambient ozone
exposures in studies not restricted to children with asthma.
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Table 6-11
Study
Linn et al. (1996)
Hoppe et al. (2003)
Scarlett etal. (1996)
Chen etal. (1999)
Avoletal. (1998a)
Studies of children
Ulmer etal. (1997)
Ward etal. (2002)
Gold et al. (1999)
Additional characteristics and quantitative data for studies
represented in Figure 6-8 and results from other studies in children
Location/
Population
3 southern CA
communities
Children
Munich, Germany
Children
Surrey, England
Children
3 Taiwan communities
Children
3 southern CA
communities
Children
O3Lag
0
0
1
1
0 (personal)
Os Averaging
Time
1 -h avg
30-min max (1 :00
p.m.-4:00 p.m.)
8-h max
1-hmax
24-h avg
Parameter
I ntraday change FEVi
Intraday change FVC
FEV,
FVC
FEV,
FEV,
Intraday change FEV,
Intraday change FVC
Effect Estimate (95% Cl)a
-0.56 (-0.99, -0.1 2)
-0.21 (-0.62, 0.20)
-1 .4 (-4.3, 1 .4)
-2.5 (-4.9, -0.10)
-0.04 (-0.32, 0.23)
-1.5 (-2.8, -0.12)
-1 .4 (-3.8, 0.90)b
-2.0 (-4.0, 0.01)b
not included in Figure 6-8°
Freudenstadt and
Villingen, Germany
Children
Birmingham and
Sandwell, England
Children
Mexico City, Mexico
Children
1
0
0-6 avg
1
1-10 avg
1/2-hmax
24-h avg
24-h avg
FEV, (ml)
PEF (L/min)
Intraday change PEF (%
change)
-5.9 (-10.4,1.3)"
-3.2 (-8.3, 2.0)d
-11.1 (-22.0, -0.18)d
-0.54 (-1.1, 0.05)
aEffect estimates are standardized to a 40-, 30-, and 20-ppb increase for 1-h (or 30-min) max, 8-h max, and 24-h avg O3,
respectively.
bThe 95% Cl was constructed using a standard error that was estimated from the p-value.
°Results are not presented in Figure 6-8 because sufficient data were not provided to calculate percent change in lung function or
PEF was analyzed.dEffect estimates are from analyses restricted to summer months.
1 Among the studies of children, the magnitudes of decrease in lung function per
2 standardized increment in ambient O3 exposure1 ranged from less than 1 to 4%, a range
3 similar to that estimated in children with asthma. However, in contrast with studies of
4 children with asthma, studies of children in the general population did not consistently
5 find that O3-associated decreases in lung function were accompanied by increases in
6 respiratory symptoms. Gold et al. (1999) found that lag 1 of O3 exposure was associated
7 with both decreases in PEF and increases in phlegm; however, the increase in phlegm
8 was associated with O3 exposure lagged one day whereas the PEF decrement was driven
9 by exposures lagged 2 to 4 days. Ozone was weakly associated with cough and shortness
10 of breath among children in England (Ward et al.. 2002). and O3 was associated with a
11 decrease in respiratory symptom score among children in California (Linn etal.. 1996).
12 These findings indicate that while the magnitudes of O3-associated decrease in lung
13 function may be similar in children with and without asthma, because of the higher
14 overall lung function in healthy children, the decrements may not be large enough to be
15 clinically significant in healthy children.
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
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Adults
1 In the small body of studies conducted in adults, O3 has been associated with decrements
2 in lung function in both healthy adults and those with comorbid factors (Table 6-12). In a
3 cohort of mostly healthy women, ages 19-43 years, followed for one summer season,
4 Naeher et al. (1999) observed associations between 8-h max ambient O3 exposure and
5 decreases in PEF. In a large cross-sectional study of 2,380 healthy adults (75th percentile
6 of age: 52 years) in Tel Aviv, Israel, across several lags of exposure (single day lags 0-7
7 and 0-6 avg), O3 was associated mostly with increases in FEVi, FVC, and FEVi/FVC
8 (Steinvil et al.. 2009). Another large cross-sectional study was conducted in 2,102
9 children and adults (mean age: 45 years) living near a petrochemical plant in Ulsan,
10 Korea (Son et al.. 2010). Multiple O3 exposure metrics, including concentrations
11 averaged across 13 city monitors, concentrations from the nearest monitor, inverse
12 distance-weighted concentrations, and estimates from kriging, were associated with
13 decrements in lung function; however, no particular metric consistently showed a larger
14 effect across the various lags of O3 exposure examined. Lag 0-2 avg of 8-h max O3
15 exposure was associated with the largest decrements in percent predicted FEVi (1.4-point
16 decrease [95% CI: -2.7, -0.08] per 30-ppb increase in the 8-h max of lag 0-2 avg O3
17 averaged across all monitors). Although the health status of subjects was not reported, the
18 mean percent predicted FEVi in the study population was 82.85%, indicating a large
19 proportion of subjects with underlying airway obstruction. Results from this study were
20 not adjusted for meteorological factors and thus, confounding cannot be ruled out.
21 As described in Section 6.2.1.1, controlled human exposure studies have not consistently
22 found O3-induced decreases in lung function in older adults. In an earlier study of adults
23 ages 69-95 years, Hoppe et al. (2003) did not find ambient O3 exposure-associated
24 decreases in lung function. However, recently, the Normative Aging Study found that
25 ambient O3 exposure was associated with decrements in FEVi and FVC in a group of
26 older men (Alexeeff et al.. 2008). This study in the Greater Boston area conducted
27 spirometry once every 3 years for 10 years in 900 older men (mean [SD] age = 68.9 [7.2]
28 years), most of whom were white and healthy. Among all subjects, several lags of 24-h
29 avg O3 exposure (1- to 7-day avg) were associated with decreases in FEVi (Alexeeff et
30 al.. 2008). Additionally, larger effects were estimated in adults with elevated BMI (> 30),
31 airway hyperresponsiveness, and reduced activity in antioxidant enzymes (i.e., GSTP1
32 He/Val or Val/Val variant) (Alexeeff etal.. 2008: Alexeeff etal.. 2007) (Table 6-12).
33 Larger O3-related decrements in FEVi and FVC also were observed in subjects with long
34 GT dinucleotide repeats in the promoter region of the antioxidant enzyme heme
35 oxygenase-1 (Alexeeff et al.. 2008). which has been associated with reduced inducibility
36 (Hiltermann et al.. 1998). The largest O3-related percentages of decrease in lung function
37 were observed in the group of men with airway hyperresponsiveness and elevated BMI (-
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1
2
3
4
5
5.3% FEVj [95% CI: -8.3, -2.4] per 20-ppb increase in lag 0-1 avg of 24-h avg O3). In
this cohort, O3 also was associated with decreases in lung function in adults without
airway hyperresponsiveness and BMI < 30, indicating the effects of O3 on lung function
in older adults extends to healthy older adults. However, importantly, the findings may be
generalizable only to older white men.
Table 6-12 Associations between ambient ozone exposure and changes in
lung function in studies of adults
Study
Son et al.
(2010)
Steinviletal.
(2009)
Naeheretal.
(1999)
Hoppe et al.
(2003)
Alexeeffetal.
(2008)
Alexeefetal.
(2007)
Location/ Population
Ulsan, Korea
Children and adults, ages 7-
97 yr
Tel Aviv, Israel
Healthy adults, mean age 43
yr, 75llr%-ile: 52 yr
Vinton, VA
Healthy women, ages 19-43
yr
Munich, Germany
Older adults, ages 69-95 yr
Greater Boston, MA
Older adults, mean (SD) age:
68.8 yr (7.3)
Greater Boston, MA
Older adults, mean (SD) age:
68.8 yr (7.3)
O3 Lag
0-2 avg
0
0-6 avg
0
0-4 avg
0
1
0-1 avg
0-1 avg
?3 •
Averaging
Time
8-h max
8-h avg
(10:00a.m-
6:00 p.m.)
24-h avg
30-min max
(1:00p.m.-
4:00p.m.)
24-h avg
24-h avg
Parameter
Change in %
predicted FEV,
FEV, (ml)
Evening PEF
(L/min)
% change in
evening FEV,
% change in FEVi
% change in FEVi
Os Assessment
Method/Subgroup
All monitor avg
Nearest monitor
IDW
Kriging
GSTP1 lie/lie
GSTP1 NeA/al ValA/al
BMI < 30
BMI > 30
NoAHR
AHR
BMI > 30 and AHR
Effect Estimate
(95% Cl)a
-1.4 (-2.7, -0.08)
-0.76 (-1 .8, 0.25)
-1.1 (-2.2,0.05)
-1.4 (-2.6, -0.11)
40 (0, 80)
94(33,156)
-0.06 (-0.11,0)
-5.1 (-8.7, -1.5)
0.75 (-2. 1,3.7)
1 .2 (-1 .3, 3.6)
-1.0 (-2.2, 0.1 9)
-2.3 (-3.5, -1.0)
-1.5 (-2.5, -0.52)
-3.5 (-5.1, -1.9)
-1.7 (-2.7, -0.73)
-4.0 (-6.2, -1.8)
-5.3 (-8.2, -2.3)
IDW = Inverse distance weighting, BMI = Body mass index, AHR = airway hyperresponsiveness.
'Effect estimates are standardized to a 40-ppb increase for 30-min max 03, 30-ppb increase for 8-h max or 8-h avg 03, and 20-ppb increase for 24-
h avg 03.
6
1
8
9
10
11
12
13
14
15
16
17
18
Confounding in epidemiologic studies of lung function
The 1996 O3 AQCD noted uncertainty regarding confounding by temperature and pollen
(U.S. EPA. 1996a); however, studies collectively do not provide strong evidence of
confounding by these factors. Most studies, whether they involved year-round or
summer-only examinations, included temperature in statistical analyses and found
associations between O3 exposure and decreases in lung function. Across studies,
temperature has shown inconsistent associations with lung function, even among studies
conducted in the summer and in the same geographic region. For example, in studies of
children attending summer camps conducted in the Northeast U.S., temperature was
associated with an increase (Berry et al.. 1991) (Thurston et al.. 1997) and decrease
(Raizenne etal., 1987) in lung function. In the reanalysis of six camp studies,
investigators did not include temperature in models because temperature within the
normal ambient range had not been shown to affect O3-induced lung function responses
in controlled human exposure studies (Kinney et al.. 1996). In two summer camp studies
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1 conducted in the Northeast U.S., O3 was associated with decreases in lung function in
2 models without and with temperature (Thurston et al.. 1997; Spektor et al.. 1988a). In
3 both studies, temperature and O3 were measured on site of the camps. Spektor et al.
4 (1988a) estimated similar effects in a model with and without a temperature-humidity
5 index, and Thurston et al. (1997) found that compared with a univariate model, O3 was
6 associated with a nearly 2-fold greater decrease in PEF when temperature was added to
7 the model.
8 Although evaluated in fewer studies, the evidence does not indicate that associations
9 between ambient O3 exposure and lung function are confounded by pollen. Some camp
10 studies found that pollen independently was not associated with lung function decrements
11 (Thurston et al.. 1997; Avol et al.. 1990). A few studies of children with asthma with
12 follow-up over multiple seasons found O3 to be associated with decrements in lung
13 function in models that adjusted for pollen counts (Just et al.. 2002; Ross et al.. 2002;
14 Jalaludin et al.. 2000; Gielen et al., 1997). In these studies, large percentages of subjects
15 had positive atopy (22-98%), with some studies examining large percentages of subjects
16 specifically with pollen allergy(Ross et al.. 2002; Gielen et al.. 1997).
17 A relatively larger number of studies provided information on potential confounding by
18 copollutants such as PM25, PM10, NO2, or SO2. In most cases, investigators indicated that
19 associations between O3 exposure and lung function were not driven by copollutant
20 confounding; however, studies varied in how they considered confounding. Studies of
21 subjects exercising outdoors indicated that ambient concentrations of copollutants such as
22 NO2, sulfur dioxide, or acid aerosol were low and thus, not likely to confound the
23 observed O3 effects (Hoppe et al.. 2003; Brunekreef etal.. 1994; Hoeketal.. 1993). In
24 other studies of children with increased outdoor exposures, O3 was consistently
25 associated with decreases in lung function, whereas other pollutants such as PM2 5,
26 sulfate, and acid aerosol individually showed variable associations across studies
27 (Thurston et al.. 1997; Castilleios et al.. 1995; Berry etal.. 1991; Avol etal.. 1990;
28 Spektor et al.. 1988a).
29 Among studies that conducted copollutant modeling, associations between O3 exposure
30 and lung function decrements were observed to be robust (Figure 6-9 and Table 6-13). In
31 copollutant models, O3 effect estimates generally fell within the 95% CI of the single-
32 pollutant model effect estimates. Whereas some studies used the same averaging time for
33 all pollutants (Lewis etal.. 2005; Jalaludin et al.. 2000). most examined 1-h max or 8-h
34 max O3 exposures and 24-h avg copollutant exposures (Son et al.. 2010; Chen et al..
35 1999; Romieuetal.. 1997; Romieuetal.. 1996). In a Philadelphia-area summer camp
36 study, Neas et al. (1999) was among the few studies to find that the effect of O3 was
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1 attenuated in a copollutant model. In a copollutant model with 24-h avg sulfate, the 12-h
2 avg O3 effect estimate was attenuated to near zero (Figure 6-9 and Table 6-13).
3 In studies with copollutant modeling, ambient O3 concentrations showed a wide range of
4 correlations with concentrations of copollutants (r=-0.31 to 0.74). Among children with
5 asthma in Sydney, Australia, Jalaludin et al. (2000) found low correlations of 24-h avg O3
6 with 24-h avg PM10 (r = 0.13) and NO2 (r = -0.31), and in two-pollutant models, PM10
7 and NO2 continued to be associated with increases in PEF, and O3 continued to be
8 associated with decreases in PEF. In a study of children with asthma in Detroit, MI, 24-h
9 avg O3 was moderately correlated with 24-h avg PM2 5 (Pearson r= 0.57) and 24-h avg
10 PM10 (Pearson r=0.59) (Lewis et al., 2005). Inclusion of PM10 or PM25 in models resulted
11 in larger changes in O3 effect estimates than those observed in other studies. As
12 illustrated in Figure 6-9 and Table 6-13, the magnitude of change was not consistent
13 between the two subgroups. Among subjects with a concurrent URI, O3-associated
14 decreases in lowest daily FEVi were robust to the inclusion of PM10 or PM2 5. Among CS
15 users, O3 was associated a much larger decrease in FEVi when PMi0 was included in the
16 model (Lewis et al.. 2005).
17 Studies conducted in Mexico City found small changes in O3-associated lung function
18 decrements in copollutant models, although different averaging times were used for
19 different pollutants (Romieu et al.. 1997; Romieuetal.. 1996) (Figure 6-9 and Table 6-
20 13). In these studies, O3 was moderately correlated with co-pollutants such as NO2 and
21 PMio (range of Pearson r = 0.38-0.58). Studies conducted in Asia also found that
22 associations between O3 and lung function were robust to the inclusion of weakly- to
23 moderately-correlated copollutants (Son et al.. 2010; Chenet al.. 1999). Copollutant
24 effect estimates generally were attenuated, indicating that O3 may confound the results of
25 copollutants.
26 In a summer camp study conducted in Connecticut, Thurston et al. (1997) found ambient
27 concentrations of 1-h max O3 and 12-h avg sulfate to be highly correlated (r = 0.74),
28 making it more difficult to separate their independent effects. With sulfate in the model, a
29 larger decrease in PEF was estimated for O3; however, the 95% CI was much wider
30 (Figure 6-9 and Table 6-13). Investigators found that the association between sulfate and
31 PEF was driven by one day when the ambient concentrations of both pollutants were at
32 their peak. With the removeal of this influential day, the sulfate effect was attenuated,
33 whereas O3 effects remained robust (Thurston et al.. 1997). Among children with asthma
34 in Thailand, the O3-associated decrease in PEF was robust to the adjustment of SO2;
35 however, different lags were examined for O3 (lag 5) and SO2 (lag 4) (Wiwatanadate and
36 Trakultivakorn. 2010). Some studies did not provide quantitative results but reported that
37 O3 effects on lung function decrements remained statistically significant in models that
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1
2
included copollutants such as PM10, NO2, sulfate, nitrate, or ammonium (Romieu et al.
1998a: Braueretal.. 1996: Linnetal.. 1996: Spektor et al.. 1988b).
Study
PEF
Jalaludin etal. (2000)
Population
O3 exposure data O3 with copollutant
24-h avg, Lag 0 None
24-havg, Lag 0 24-h avg PM10
24-h avg, Lag 0 24-h avg NO2
Neasetal. (1999)= Children attendingcamp 12-havg,Lag1 None
1 2-h avg, Lag 1 24-h avg su Ifate
Thurston eta,. (1997)
^ «9 ° None
1-h max, Lag 0 12-h avg sulfate-4-
Romieu etal. (1996) Children with asthma 1-h max, Lag 0 None
1-h max, Lag 0 24-h avg PM2 5
Romieu etal. (1997) Children with asthma 1-h max, Lag 0 None
1-h max, Lag 0 24-h avg PM10
tant
-O
-O
C
— •-
)
-
-15 -13 -11 -9 -7 -5 -3-11 3
Percent change in PEF per standardized increase in O3 (95% Cl)
FEV,
Lewis etal. (2005)
Chen etal. (1999)
Children with asthrna 24_h avg] Lag 2
24-h avg, Lag 2
Children with asthma
withURI 24-h avg, Lag 2
24-h avg, Lag 2
24-h avg, Lag 2
Children 1-h max, Lag 1
1-h max, Lag 1
id.g . ,„
'4ha"i'rM"
24-h avg NO2 0
-15 -13 -11 -9 -7 -5 -3-11 3
Percent change in FEV, per standardized increase in O3 (95% Cl)
Results are presented for PEF then FEV! and then in order of increasing mean ambient ozone concentration. "Information was not
available to calculate 95% Cl of the copollutant model. CS = corticosteroid, URI = Upper respiratory infection. Effect estimates are
standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 12-h avg, and 24-h avg ozone, respectively. Black circles represent
ozone effect estimates from single pollutant models, and open circles represent ozone effect estimates from copollutant models.
Figure 6-9 Comparison of ozone-associated changes in lung function in
single- and copollutant models.
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Table 6-13 Additional characteristics and quantitative data for studies
presented in Figure 6-9
Study
Location/
Population
Os Exposure
Data
Parameter
Os-associated
Percent Change in
Single-Pollutant
Model (95% Cl)a
Os-associated Percent
Change in Copollutant
Model (95% Cl)a
PEF
Jalaludin et al.
(2000)
Neas et al. (1999)
Thurston et al.
(1997)
Romieu et al.
(1996)
Romieu et al.
(1997)
Sydney, Australia
Children with asthma or
wheeze
Philadelphia, PA
Children attending summer
camp
CT River Valley
Children with asthma attending
summer camp
Mexico City, Mexico
Children with asthma
Mexico City, Mexico
Children with asthma
24-h avg
LagO
12-havg
Lag1
1-hmax
LagO
1-hmax
LagO
1-h max
LagO
Intraday change
PEF
Morning PEF
Intraday change
PEF
Evening PEF
Evening PEF
-0.57 (-1.1, -0.06)
-0.94 (-2.0, 0.08)
-2.8 (-4.9, -0.59)
-0.55 (-1.3, 0.1 9)
-0.52 (-1.0, -0.01)
with 24-h avg PM10,
-0.57 (-1.1, -0.06)
with 24-h avg N02
-0.55 (-0.1, -0.04)
with 24-h avg sulfate
-0.02b
with 12-havg sulfate
-11. 8 (-31 .6, 8.1)
with 24-h avg PM2.5
-0.24 (-1.2, 0.68)
with 24-h avg PM10
-0.79 (-1.4, -0.16)
FEV,
Lewis etal.
(2005)
Chen etal. (1999)
Detroit, Ml
Children with asthma using
CS
Children with asthma with
URI
3 Taiwan communities
Children
24-h avg
Lag 2
1-hmax
Lag1
Lowest daily
FEV,
FEV,
0.29 (-4.2, 5.0)
-6.0 (-11. 2, -0.41)
-1.5 (-2.8, -0.12)
with 24-h avg PM2.5
-0.1 8 (-11. 0,11.9)
with 24-h avg PM10
-13.4 (-17.8, -8.8)
with 24-h avg PM2.5
-5.5 (-10.3, -0.42)
with 24-h avg PM10
-7.1 (-11.3, -2.8)
with 24-h avg N02
-2.0 (-3.5, 0.42)
Results not included In Figure 6-9
Wiwatanadate
and
Trakultivakorn
(2010)
Son et al. (2010)
Chiang Mai, Thailand
Children with asthma
Ulsan, Korea
Children and adults
24-h avg
Lag5
8-h max
Lag 0-2 avg
Evening PEF
(L/min)
Change in %
predicted FEVi
-2.6 (-5.2, 0)
-1.4 (-2.6, -0.11)
with Lag 4 S02
-3.2 (-6.2, -0.2)
with PM10
-1.8 (-3.4, -0.25)
(kriging)
CS = Corticosteroid, URI = Upper respiratory infection.
aResults represent percent changes in lung function parameter per the following standardized increase in ambient O3
concentration: 40 ppb for 1-h max O3, 30 ppb for 8-h max or 12-h avg O3, and 20 ppb for 24-h avg O3.
1 Several studies examined multi-pollutant models that most often included O3, NO2, and
2 either PM2 5 or PM10. Ozone exposure was associated with similar or larger magnitudes of
3 decrease lung function in multi-pollutant models (O'Connor et al.. 2008; Thaller et al..
4 2008; Chan and Wu. 2005; Romieu et al.. 2002: Korrick et al.. 1998: Higgins et al..
5 1990): however, the independent effects of O3 exposure are more difficult to assess in
6 relation to incremental changes in more than one copollutant.
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Summary of Epidemiologic Studies of Lung Function
1 The cumulative body of epidemiologic evidence strongly supports associations between
2 ambient O3 exposure and decrements in lung function in children, particularly, those with
3 asthma. While little new research is available, previous AQCDs have presented
4 epidemiologic evidence of heightened effects in children and adults exercising or
5 working outdoors during periods of relatively low ambient O3 concentrations (Table 6-1).
6 These epidemiologic results are well-supported by observations from controlled human
7 exposure studies in which exposures to lower O3 concentrations induce lung function
8 decrements when combined with exercise as compared with exposures during rest.
9 Recent epidemiologic investigation continued to focus on children with asthma, and most
10 recent results in this population indicated associations between O3 exposure and
11 decrements in lung function (Figures 6-6 anf 6-7 and Tables 6-7 and 6-8). Based on a
12 small number of within-study comparisons of groups with and without asthma, larger
13 effects were not conclusively estimated for groups with asthma. It is important to note
14 that most of these studies were not designed to assess between-group differences, and in
15 some studies, the high prevalence of atopy may have contributed to larger associations in
16 subjects without asthma (Khatri et al.. 2009; Barraza-Villarreal et al.. 2008). A large body
17 of previous studies demonstrated associations in children. Whereas the 2006 O3 AQCD
18 reported weak evidence, a new study indicates that O3 exposure may be associated with
19 decrements in lung function in older adults.
20 Across the diverse populations examined in epidemiologic studies, ambient O3 exposure
21 was associated with 1-8% decreases in lung function per standardized increment in O3
22 concentration1. Larger decreases (3-8%) usually were observed in children with asthma
23 or older adults with CS use, concurrent URI, airway hyperresponsiveness, or reduced
24 activity of antioxidant enzymes. These results indicate that common comorbid and
25 genetic factors may increase the risk of O3-associated respiratory effects. High dietary
26 antioxidant intake was found to attenuate O3-associated lung function decrements. Each
27 of these potential susceptibility or protective factors has been examined in one to two
28 populations, and further investigation in diverse populations is warranted. Heterogeneity
29 in response also was demonstrated by observations that increases in ambient O3 exposure
30 were associated with increased incidence of a greater than 10% decline in lung function
31 in children with asthma (Hoppe et al.. 2003; Mortimer et al.. 2002). In considering the
32 clinical significance of more subtle health outcomes such as lung function changes, it is
33 important to note that a small shift in the population mean likely will have a
34 disproportionate effect in the extreme ends of the distribution of lung function where
35 these small magnitudes of decrease lead to clinically-significant airway resistance or
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
Draft - Do Not Cite or Quote 6-54 September 2011
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1 obstruction and where individuals likely have concurrent symptoms. Several
2 epidemiologic studies have demonstrated the clinical significance of O3-associated lung
3 function decrements, primarily in individuals with asthma, by finding concomitant
4 increases in respiratory symptoms (Khatri et al.. 2009; Just et al.. 2002; Mortimer et al..
5 2002; Ross et al.. 2002; Gielenetal.. 1997; Romieuetal.. 1997; Thurston et al.. 1997;
6 Romieuetal.. 1996).
7 Collectively, epidemiologic studies have examined and found decreases in lung function
8 in association with single-day O3 concentrations lagged from 0 to 7 days as well
9 concentrations averaged over 2-10 days. A large body of evidence indicates decreases in
10 lung function in association with O3 exposures over the duration of outdoor activity,
11 same-day, or previous-day O3 exposures (Sonet al.. 2010; Alexeeff et al.. 2008; Lewis et
12 al.. 2005; Ross et al.. 2002; Jalaludin et al.. 2000; Chen et al.. 1999; Romieuetal.. 1997;
13 Braueretal. 1996: Romieuetal.. 1996: Spektor et al.. 1988b). Fewer studies find
14 associations with longer lags of ambient O3 exposures (5-7 days) (Wiwatanadate and
15 Trakultivakorn. 2010: Hernandez-Cadena et al.. 2009: Steinvil et al.. 2009). However,
16 associations with multiday averages of exposure (Sonet al.. 2010: Liu et al.. 2009a:
17 Barraza-Villarreal et al.. 2008: O'Connor et al.. 2008: Alexeeff et al.. 2007: Mortimer et
18 al.. 2002: Ward et al.. 2002: Gold et al.. 1999: Naeher et al.. 1999: Neas et al.. 1999)
19 indicate that exposures accumulated over several days may be important. For single- and
20 multi-day O3 exposures, associations with lung function decrements were observed for 1-
21 h max, 8-h max, and 24-h avg O3, without a clear indication that the strength of evidence
22 varied among the averaging times. Within studies, O3 exposure for various lag periods
23 were associated with lung function decrements, possibly indicating that multiple modes
24 of action may be involved in the responses. Activation of bronchial C-fibers (Section
25 5.3.2) may lead to decreases in lung function as an immediate response to O3 exposure,
26 and increased airway hyperresponsiveness resulting from sensitization of airways
27 (Section 5.3.5) may mediate lung function responses associated with the lagged or
28 multiday O3 exposures (Peden. 2011).
29 Several studies found that associations with lung function decrements persisted at lower
30 ambient O3 concentrations. For exposures averaged up to 1 hour during outdoor activity,
31 multiple studies in individuals engaged in outdoor activities found associations with O3
32 concentrations limited to those below 80 ppb (Spektor etal., 1988a: Spektor et al..
33 1988b). 60 ppb (Brunekreef et al.. 1994: Spektor et al.. 1988a). and 50 ppb (Brunekreef et
34 al.. 1994). Among outdoor workers, Brauer et al. (1996) found a robust association with
35 daily 1-h max O3 concentrations below 40 ppb. For 8-h average O3 exposures,
36 associations with lung function decrements in children with asthma were found to persist
37 at concentrations less than 80 ppb in a U.S. multicity study (for lag 1-5 avg) (Mortimer et
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1 al.. 2002) and less than 51 ppb in a study conducted in the Netherlands (for lag 2) (Gielen
2 etal.. 1997).
3 Several studies of lung function evaluated confounding by meterological factors and
4 copollutant exposures. Most O3 effect estimates remained robust in models that adjusted
5 for temperature, humidity, and co-pollutants such as PM2s, PMio, NO2, or SO2. Although
6 examined in relatively few epidemiologic studies, O3 was associated with decreases in
7 lung function in models that included pollen or acid aerosols. The consistency of
8 association in the collective body of evidence with and without adjustment for
9 copollutant exposures and meterological factors combined with evidence from controlled
10 human exposure studies for the direct effect of O3 exposure provide substantial evidence
11 for the independent effects of ambient O3 exposure on lung function decrements.
6.2.1.3 Toxicology
12 The 2006 O3 AQCD concluded that pulmonary function decrements occur in a number of
13 species with acute exposures (< 1 week), ranging from 0.25 to 0.4 ppm O3 (U.S. EPA.
14 2006b). Early work has demonstrated that during acute exposure of ~0.2 ppm O3 in rats,
15 the most commonly observed alterations are increased frequency of breathing and
16 decreased tidal volume (i.e., rapid, shallow breathing). Decreased lung volumes are
17 observed in rats with acute exposures to 0.5 ppm O3. At concentrations of >1 ppm,
18 breathing mechanics (compliance and resistance) are also affected. Exposures of 6 h/day
19 for 5 days create a pattern of attenuation of pulmonary function decrements in both rats
20 and humans without concurrent attenuation of lung injury and morphological changes,
21 indicating that the attenuation did not result in protection against all the effects of O3
22 (Wiester et al.. 1996b). A number of studies examining the effects of O3 on pulmonary
23 function in rats, mice, and dogs are described in Table 6-13 on p. 6-91 of the 1996 O3
24 AQCD and Table AX5-11 on p. AX5-34 of the 2006 O3 AQCD (U.S. EPA. 2006b.
25 1996a). Recent lung imaging studies using hyperpolarized 3He provide evidence of
26 ventilation abnormalities in rats following exposure to 0.5 ppm O3 (Cremillieux et al..
27 2008). Rats were exposed to 0.5 ppm O3 for 2 or 6 days, either continuously (22 h/day) or
28 alternating (12 h/day). Dynamic imaging of lung filling (2 mL/s) revealed delayed and
29 incomplete filling of lung segments and lobes. Abnormalities were mainly found in the
30 upper regions of the lungs and proposed due to the spatial distribution of O3 exposure
31 within the lung. Although the small number of animals used in the study (n = 3 to
32 7/group) makes definitive conclusions difficult, the authors suggest that the delayed
33 filling of lung lobes or segments is likely a result of an increase in airway resistance
34 brought about by narrowing of the peripheral small airways.
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6.2.2 Airway Hyperresponsiveness
1 Airway hyperresponsiveness refers to a condition in which the conducting airways
2 undergo enhanced bronchoconstriction in response to a variety of stimuli. Airway
3 responsiveness is typically quantified by measuring changes in pulmonary function (e.g.,
4 FEVi or specific airway resistance [sRaw]) following the inhalation of an aerosolized
5 specific (allergen) or nonspecific (e.g., methacholine) bronchoconstricting agent or
6 another stimulus such as exercise or cold air. Asthmatics are generally more sensitive to
7 bronchoconstricting agents than nonasthmatics, and the use of an airway challenge to
8 inhaled bronchoconstricting agents is a diagnostic test in asthma. Standards for airway
9 responsiveness testing have been developed for the clinical laboratory (ATS. 2000aX
10 although variation in methodology for administering the bronchoconstricting agent may
11 affect the results (Cockcroft et al.. 2005). There is a wide range of airway responsiveness
12 in nonasthmatic people, and responsiveness is influenced by wide range of factors,
13 including cigarette smoke, pollutant exposures, respiratory infections, occupational
14 exposures, and respiratory irritants. Airways hyperresponsiveness in response to O3
15 exposure has not been examined widely in epidemiologic studies; such evidence is
16 derived primarily from controlled human exposure and toxicological studies.
6.2.2.1 Controlled Human Exposures
17 Beyond its direct effect on lung function, O3 exposure causes an increase in airway
18 responsiveness in human subjects as indicated by a reduction in the concentration of
19 specific (e.g., ragweed) and non-specific (e.g., methacholine) agents required to produce
20 a given reduction in FEVi or increase in sRaw. Increased airway responsiveness is an
21 important consequence of exposure to ambient O3, because the airways are then
22 predisposed to narrowing upon inhalation of a variety of ambient stimuli including
23 specific allergens, SO2, and cold air.
24 Increases in airway responsiveness have been reported for exposures to 80 ppb O3 and
25 above. Horstman et al. (1990) evaluated airway responsiveness to methacholine in young
26 healthy adults (22 M) exposed to 80, 100, and 120 ppb O3 (6.6 h, quasi continuous
27 moderate exercise, 39 L/min). Dose-dependent decreases of 33, 47, and 55% in the
28 cumulative dose of methacholine required to produce a 100% increase in sRaw after
29 exposure to O3 at 80, 100, and 120 ppb, respectively, were reported. Molfino et al. (1991)
30 reported increased allergen-specific airway responsiveness in mild asthmatics exposed to
31 120 ppb O3 (1 h resting exposure). Due to safety concerns, however, the exposures in the
32 Molfino et al. (1991) study were not randomized with FA conducted first and O3
33 exposure second. Attempts to reproduce the findings of Molfino et al. (1991) using a
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1 randomized exposure design have not found statistically significant changes in airway
2 responsiveness at such low levels of O3 exposure. At a considerably higher exposure to
3 250 ppb O3 (3 h, light-to-moderate intermittent exercise, 30 L/min), Torres et al. (1996)
4 found significant increases in specific and non-specific airway responsiveness of mild
5 asthmatics 3 h following O3 exposure. Kehrl et al. (1999) found increased reactivity to
6 house dust mite antigen in mild atopic asthmatics 16-18 h after exposure to 160 ppb O3
7 (7.6 h, light quasi continuous exercise, 25 L/min). Holz et al. (2002) demonstrating that
8 repeated daily exposure to lower concentrations of 125 ppb O3 (3 h for four consecutive
9 days; intermittent exercise, 30 L/min) causes an increased response to allergen challenge
10 at 20 h postexposure in allergic airway disease.
11 O3 exposure of asthmatic subjects, who characteristically have increased airway
12 responsiveness at baseline relative to healthy controls (by nearly two orders of
13 magnitude), can cause further increases in responsiveness (Kreit et al.. 1989). Similar
14 relative changes in airway responsiveness are seen in asthmatics and healthy control
15 subject exposed to O3 despite their markedly different baseline airway responsiveness.
16 Several studies (Kehrl et al.. 1999; Torres et al.. 1996; Molfino et al.. 1991) have
17 suggested an increase in specific (i.e., allergen-induced) airway reactivity. An important
18 aspect of increased airway responsiveness after O3 exposure is that this may represent a
19 plausible link between ambient O3 exposure and increased respiratory symptoms in
20 asthmatics, and increased hospital admissions and ED visits for asthma.
21 Changes in airway responsiveness after O3 exposure appear to resolve more slowly than
22 changes in FEVi or respiratory symptoms (Folinsbee and Hazucha, 2000). Studies
23 suggest that O3-induced increases in airway responsiveness usually resolve 18 to 24 h
24 after exposure, but may persist in some individuals for longer periods (Folinsbee and
25 Hazucha. 1989). Furthermore, in studies of repeated exposure to O3, changes in airway
26 responsiveness tend to be somewhat less susceptible to attenuation with consecutive
27 exposures than changes in FEVi (Gong et al.. 1997a: Folinsbee et al.. 1994; Kulle et al..
28 1982; Dimeo etal., 1981). Increases in airway responsiveness do not appear to be
29 strongly associated with decrements in lung function or increases in symptoms (Aris et
30 al., 1995). Recently, Que et al. assessed methacholine responsiveness in healthy young
31 adults (83M, 55 F) at one day after exposure to 220 ppb O3 and FA for 2.25 h (alternating
32 15 min periods of rest and brisk treadmill walking). Increases in airways responsiveness
33 at 1 day post-O3 exposure were not correlated with FEVi responses immediately
34 following the O3 exposure nor with changes in epithelial permeability assessed 1 day
35 post-O3 exposure.
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6.2.2.2 Toxicology
1 In addition to human subjects, a number of species, including nonhuman primates, dogs,
2 cats, rabbits, and rodents, have been used to examine the effect of O3 exposure on airway
3 hyperresponsiveness. With a few exceptions, commonly used animal models have been
4 guinea pigs, rats, or mice acutely exposed to O3 concentrations of 1 to 3 ppm to induce
5 airway hyperresponsiveness. These animal models are helpful for determining underlying
6 mechanisms of general airway hyperresponsiveness and are relevant for understanding
7 airway responses in humans. Although 1-3 ppm may seem like a high exposure
8 concentration, based on 18O3 (oxygen-18-labeled ozone) in the BALF of humans and rats,
9 an exposure of 0.4 ppm O3 in exercising humans appears roughly equivalent to an
10 exposure of 2 ppm in resting rats (Hatch etal.. 1994).
11 A limited number of studies have observed airway hyperresponsiveness in rodents and
12 guinea pigs after exposure to less than 0.3 ppm O3. As previously reported in the 2006 O3
13 AQCD, one study demonstrated that a very low concentration of O3 (0.05 ppm for 4 h)
14 induced airway hyperresponsiveness in some of the nine strains of rats tested (Depuydt et
15 al., 1999). This effect occurred at a concentration of O3 that was much lower than has
16 been reported to induce airway hyperresponsiveness in any other species. Similar to
17 ozone's effects on other endpoints, these observations suggest a genetic component plays
18 an important role in O3-induced airway hyperresponsiveness in this species and warrants
19 verification in other species. More recently, Chhabra and colleagues (2010) demonstrated
20 that exposure of ovalbumin (OVA)-sensitized guinea pigs to 0.12 ppm for 2 h/day for 4
21 weeks produced specific airway hyperresponsiveness to an inhaled OVA challenge.
22 Interestingly, in this study , dietary supplementation of the guinea pigs with vitamins C
23 and E ameliorated a portion of the airway hyperresponsiveness as well as indices of
24 inflammation and oxidative stress. Larsen and colleagues did an O3 concentration-
25 response study in mice sensitized by 10 daily inhalation treatments with an OVA aerosol
26 (Larsen et al.. 2010). Although airway responsiveness to methacholine was increased in
27 non-sensitized animals exposed to a single 3-h exposure to 0.5, but not 0.1 or 0.25, ppm
28 O3, airway hyperresponsiveness was observed after exposure to 0.1 and 0.25 ppm O3 in
29 OVA-sensitized mice. Shore and colleagues (Johnston et al.. 2005b) have also
30 demonstrated O3-induced airway hyperresponsiveness in mice after exposure to 0.3 ppm
31 O3 for 3 hours. Mice that were exposed to the same concentration of O3 for 72 hours
32 showed no evidence of airway hyperresponsiveness, indicating attenuation of this effect.
33 Thus, recent toxicological studies have demonstrated that O3-induced airway
34 hyperresponsiveness occurs in guinea pigs and mice after either acute or repeated
35 exposure to relevant concentrations of O3.
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1 The mechanisms by which O3 enhances the airway responsiveness to either specific (e.g.,
2 OVA) or non-specific (e.g., methacholine) bronchoprovocation are not clear, but appear
3 to be associated with complex cellular and biochemical changes in the conducting
4 airways. Considerable research effort has been directed towards exploring the causes of
5 O3-induced airway hyperresponsiveness, but the majority of such studies have been
6 conducted at high concentrations of O3. It is clear that inflammation plays a key role in
7 O3-induced airway hyperresponsiveness, although the precise mediators and cells that are
8 involved have not been identified at relevant concentrations of O3. Because inflammation
9 is likely to play a role in O3-induced airway hyperresponsiveness, the mechanism for this
10 response may be multifactorial, involving the presence of cytokines, prostanoids, or
11 neuropeptides; activation of macrophages, eosinophils, or mast cells; and epithelial
12 damage that increases direct access of mediators to the smooth muscle or receptors in the
13 airways that are responsible for reflex bronchoconstriction. Johnston et al. (2005b)
14 demonstrated that airway hyperresponsiveness occurred in both wild type and IL-6
15 knockout mice exposed to 0.3 ppm O3 despite reduction in markers of lung injury and
16 inflammation in O3-exposed IL-6 knockout mice. This same group of investigators has
17 demonstrated the involvement of natural killer T cells, obesity, CXCR2, leptin, and IL-17
18 in O3-induced airway hyperresponsiveness at exposure concentrations of 1-3 ppm O3
19 (Garantziotis et al.. 2010; Voynow et al.. 2009; Pichavant et al.. 2008; Williams et al..
20 2007b: Lu et al.. 2006; Johnston et al.. 2005a: Shore et al.. 2003). A recent study
21 demonstrated a role for mindin, an extracellular matrix protein, in the AHR response
22 resulting from acute exposure to 1 ppm O3 (Frush et al.. In Press). Thus, a number of
23 potential mediators and cells may play a role in O3-induced airway hyperresponsiveness;
24 mechanistic studies are discussed in greater detail in Chapter 5.
25 In order to evaluate the ability of O3 to enhance specific and non-specific airway
26 responsiveness, it is important to take into account the phenomenon of attenuation in
27 ozone's effects. Several studies have clearly demonstrated that some effects caused by
28 acute exposure are absent after repeated exposures to O3. The ability of the pulmonary
29 system to adapt to repeated insults to O3 is complex, however, and experimental findings
30 for attenuation to O3-induced airway hyperresponsiveness are inconsistent. As described
31 above, airway hyperresponsiveness was observed in mice after a 3-h exposure but not in
32 mice exposed continuously for 72 hours to 0.3 ppm (Johnston et al.. 2005b). However,
33 the Chhabra study demonstrated O3-induced airway hyperresponsiveness in guinea pigs
34 exposed for 2 h/day for 10 days (Chhabra et al.. 2010). Besides the obvious species
35 disparity, these studies differ in that the mice were exposed continuously for 72 hours,
36 whereas the guinea pigs were exposed intermittently over 10 days, suggesting that
37 attenuation might be lost with periods of rest in between O3 exposures.
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6.2.3 Pulmonary Inflammation, Injury and Oxidative Stress
1 In addition to physiological pulmonary responses, respiratory symptoms, and airway
2 hyperresponsiveness, O3 exposure has been shown to result in increased epithelial
3 permeability and respiratory tract inflammation. In general, inflammation can be
4 considered as the host response to injury and the induction of inflammation as evidence
5 that injury has occurred. Inflammation induced by exposure of humans to O3 can have
6 several potential outcomes: (1) inflammation induced by a single exposure (or several
7 exposures over the course of a summer) can resolve entirely; (2) continued acute
8 inflammation can evolve into a chronic inflammatory state; (3) continued inflammation
9 can alter the structure and function of other pulmonary tissue, leading to diseases such as
10 fibrosis; (4) inflammation can alter the body's host defense response to inhaled
11 microorganisms, particularly in potentially susceptible populations such as the very
12 young and old; and (5) inflammation can alter the lung's response to other agents such as
13 allergens or toxins. Except for outcome (1), the possible chronic responses have only
14 been directly observed in animals exposed to O3. It is also possible that the profile of
15 response can be altered in persons with preexisting pulmonary disease (e.g. asthma,
16 COPD) or smokers. Oxidative stress has been shown to play a key role in initiating and
17 sustaining O3-induced inflammation. Secondary oxidation products formed as a result of
18 reactions between O3 and components of the ELF can increase the expression of
19 cytokines, chemokines, and adhesion molecules and enhance airway epithelium
20 permeability (Sections 5.3.3. and 5.3.4.).
6.2.3.1 Controlled Human Exposures
21 As reported in studies reviewed in the 1996 and 2006 O3 AQCDs, acute O3 exposure
22 initiates an acute inflammatory response throughout the respiratory tract which has been
23 observed to persist for at least 18-24 hours postexposure. A meta-analysis of 21 studies
24 (Mudway and Kelly. 2004a) showed that neutrophils (PMN) influx in healthy subjects
25 was linearly associated (p<0.01) with total O3 dose (i.e., the product of O3 concentration,
26 exposure duration, and VE). As with FEVi responses to O3, within individual
27 inflammatory responses to O3 are generally reproducible and correlated between repeat
28 exposures (Holz et al.. 1999). Some individuals also appear to be intrinsically more
29 susceptible to increased inflammatory responses to O3 exposure (Holz et al.. 2005).
30 The presence of PMNs in the lung has long been accepted as a hallmark of inflammation
31 and is an important indicator that O3 causes inflammation in the lungs. Neutrophilic
32 inflammation of tissues indicates activation of the innate immune system and requires a
33 complex series of events which are normally followed by processes that clear the
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1 evidence of acute inflammation. Inflammatory effects have been assessed in vivo by
2 lavage (proximal airway and bronchoalveolar), bronchial biopsy, and more recently,
3 induced sputum. A single acute exposure (1-4 hours) of humans to moderate
4 concentrations of O3 (0.2-0.6 ppm) while exercising at moderate to heavy intensities
5 results in a number of cellular and biochemical changes in the lung, including an
6 inflammatory response characterized by increased numbers of PMNs, increased
7 permeability of the epithelial lining of the respiratory tract, cell damage, and production
8 of proinflammatory cytokines and prostaglandins (U.S. EPA. 2006b). These changes also
9 occur in humans exposed to 80 and 100 ppb O3 for 6-8 hours (Alexis et al., 2010; Peden
10 et al.. 1997; Devlin et al.. 1991). Soluble mediators of inflammation such as the cytokines
11 (e.g., IL-6, IL-8) and arachidonic acid metabolites (e.g., prostaglandin [PG]E2, PGF2a,
12 thromboxane, and leukotrienes [LTs] such as LTB4) have been measured in the BALF of
13 humans exposed to O3. In addition to their role in inflammation, many of these
14 compounds have bronchoconstrictive properties and may be involved in increased airway
15 responsiveness following O3 exposure. The possible relationship between repetitive bouts
16 of acute inflammation in humans caused by O3 and the development of chronic
17 respiratory disease is unknown.
18 Studies reviewed in the 2006 O3 AQCD reported that inflammatory responses do not
19 appear to be correlated with lung function responses in either asthmatic or healthy
20 subjects. In healthy adults (14 M, 6 F) and asthmatic (12 M, 6 F) volunteers exposed to
21 200 ppb O3 (4 h with moderate quasi continuous exercise, VE = 44 L/min), percent PMN
22 and total protein in BAL fluids were significantly increased in the asthmatics relative to
23 the healthy controls. Spirometric measures of lung function were significantly decreased
24 following the O3 exposure in both groups, but were not significantly different between
25 the asthmatic and healthy subjects. Effects of O3 on PMN and total protein were not
26 correlated with changes in FEVi or FVC (Balmes et al.. 1997: Balmes et al.. 1996V
27 Devlin et al. (1991) exposed healthy adults (18 M) to 80 and 100 ppb O3 (6.6 h with
28 moderate quasi continuous exercise, 40 L/min). In BAL fluid collected 18 h after
29 exposure to 100 ppb O3, significant increases in PMNs, protein, PGE2, fibronectin, IL-6,
30 lactate dehydrogenase, and a-1 antitrypsin compared to FA. Similar but smaller increases
31 in all mediators were found after exposure to 80 ppb O3 except for protein and
32 fibronectin. Changes in BAL markers were not correlated with changes in FEVi. Holz et
33 al. (1999) examined inflammatory responses in healthy (n=21) and asthmatic (n=15)
34 subjects exposed to 125 and 250 ppb O3 (3 h, light intermittent exercise, 26 L/min).
35 Significantly increased percent PMN in sputum due to O3 exposure was observed in both
36 asthmatics and healthy subjects following the 250 ppb exposure. At the lower, 125 ppb
37 exposure, only the asthmatic group experienced statistically significantly increases in the
38 percent PMN. Significant decrements in FEVi were only found following exposure to
39 250 ppb; these changes in FEVi did not differ significantly between the asthmatic and
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1 healthy groups, nor were changes in FEVi correlated with changes in PMN levels. In
2 contrast to these earlier findings, Vagaggini et al. (2010) recently reported a significant
3 (r=0.61, p=0.015) correlation between changes in FEVi and changes in sputum
4 neutrophils in mild-to-moderate asthmatics (n=23; 33 ± 11 years) exposed to 300 ppb O3
5 for 2 hours with moderate exercise.
6 The time course of the inflammatory response to O3 in humans has not been fully
7 characterized. Different markers exhibit peak responses at different times. Studies in
8 which lavages were performed 1 hour after O3 exposure (1 h at 0.4 ppm or 4 h at
9 0.2 ppm) have demonstrated that the inflammatory responses are quickly initiated (Torres
10 etal., 1997; Devlin et al., 1996; Schelegle et al., 1991). Inflammatory mediators and
11 cytokines such as IL-8, IL-6, and PGE2 are greater at 1 h than at 18 h post-O3 exposure
12 (Torres et al.. 1997; Devlin et al., 1996). However, IL-8 still remained elevated at 18 h
13 post-O3 exposure (4 h at 0.2 ppm O3 versus FA) in healthy subjects (Balmes et al.. 1996).
14 Schelegle et al. (1991) found increased PMNs in the "proximal airway" lavage at 1, 6,
15 and 24 hours after O3 exposure (4 h at 0.2 ppm O3), with a peak response at 6 hours.
16 However, at 18-24 hours after O3 exposure, PMNs remain elevated relative to 1 hour
17 postexposure (Torres et al.. 1997; Schelegle et al.. 1991).
18 Alexis et al. (2010) recently reported that a 6.6-hour exposure with moderate exercise to
19 80 ppb O3 caused increased sputum neutrophil levels at 18 hours postexposure in young
20 healthy adults (n=15; 24 ± 1 years). In a prior study, Alexis et al. (2009) found genotype
21 effects on inflammatory responses but not lung function responses to a 2 h-exposure to
22 400 ppb O3. At 4 h post O3 exposure, both GSTM1 genotypes had significant increases in
23 sputum neutrophils with a tendency for a greater increase in GSTM1-sufficient than null
24 individuals. At 24 h postexposure, neutrophils had returned to baseline levels in the
25 GSTM1-sufficient individuals. In the GSTMl-null subjects, however, neutrophil levels
26 increased further from 4 h to 24 h and were significantly greater than both baseline levels
27 and 24 h levels in GSTM1-sufficient individuals. Alexis et al. (2009) found that GSTM1-
28 sufficient individuals (n=19; 24 ± 3 years) had a decrease in macrophage levels at 4-
29 -24 hours postexposure to 400 ppb O3 for 2 h with exercise. These studies also provide
30 evidence for activation of innate immunity and antigen presentation, as discussed in
31 Section 5.3.6. Effects of the exposure apart from O3 cannot be ruled out in the Alexis et
32 al. (2010; 2009) studies, however, since no FA exposure was conducted.
33 Kim et al. (2011) has more recently shown a significant (p < 0.001) increase in sputum
34 neutrophil levels following a 6.6-hour exposure to 60 ppb O3 relative to FA in young
35 healthy adults (13 F, 11 M; 25.0 ± 0.5 years). There was no significant effect of GSTM1
36 genotype (half GSTMl-null) on the inflammatory responses observed in these
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1 individuals. Previously, inflammatory responses had only been evaluated down to a level
2 of80ppbO3.
3 Inflammatory responses to O3 exposure have also been studied in asthmatic subjects
4 (Pedenetal.. 1997; Scannell et al.. 1996; Bashaetal.. 1994). In these studies, asthmatics
5 showed significantly more neutrophils in BALF (18 hours postexposure) than did
6 similarly exposed healthy individuals. In one of these studies (Peden et al.. 1997). which
7 included only allergic asthmatics who tested positive for Dematophagoides farinae
8 antigen, there was an eosinophilic inflammation (twofold increase), as well as
9 neutrophilic inflammation (threefold increase). In a study of subjects with intermittent
10 asthma exposed to 0.4 ppm O3 for 2 hours, increases in eosinophil cationic protein,
11 neutrophil elastase and IL-8 were found to be significantly increased 16 hours
12 postexposure and comparable in induced sputum and BALF (Hiltermann et al.. 1999).
13 Scannell et al. (1996) also reported that IL-8 tends to be higher in the BALF of
14 asthmatics compared to nonasthmatics following O3 exposure, suggesting a possible
15 mediator for the significantly increased neutrophilic inflammation in those subjects.
16 Bosson et al. (2003) found significantly greater epithelial expression of IL-5, IL-8,
17 granulocyte-macrophage colony-stimulating factor and epithelial cell-derived neutrophil -
18 activating peptide-78 in asthmatics compared to healthy subjects following exposure to
19 0.2 ppm O3 for 2 h. In contrast, Stenfors et al. (2002) did not detect a difference in the O3-
20 induced increases in neutrophil numbers between 15 mild asthmatic and 15 healthy
21 subjects by bronchial wash at the 6 h postexposure time point. However, the asthmatics
22 were on average 5 years older than the healthy subjects in this study, and it is not yet
23 known how age affects inflammatory responses. It is also possible that the time course of
24 neutrophil influx differs between healthy and asthmatic individuals. Differences between
25 asthmatics and healthy subjects in ozone-mediated activation of innate and adaptive
26 immune responses have been observed in two studies (Hernandez et al.. 2010; Bosson et
27 al., 2003). as discussed in Sections 6.2.5.4 and 5.4.2.2.
28 Vagaggini et al. (2002) investigated the effect of prior allergen challenge on responses in
29 mild asthmatics exposed for 2 h to 0.27 ppm O3 or filtered air. At 6 h postexposure,
30 eosinophil numbers in induced sputum were found to be significantly greater after O3
31 than after air exposures. Studies such as this suggest that the time course of eosinophil
32 and neutrophil influx following O3 exposure can occur at levels detectable within the
33 airway lumen by as early as 6 h. They also suggest that the previous or concurrent
34 activation of proinflammatory pathways within the airway epithelium may enhance the
35 inflammatory effects of O3. For example, in an in vitro study of primary human nasal
36 epithelial cells and BEAS-2B cell line, cytokine production induced by rhinovirus
37 infection was enhanced synergistically by concurrent exposure to O3 at 0.2 ppm for 3
38 hours (Spannhake et al.. 2002).
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1 Markers from BALF following both 2 hours (Devlin et al., 1997) and 4 hours (Torres et
2 al.. 2000; Christian et al.. 1998) repeated O3 exposures (up to 5 days) indicate that there is
3 ongoing cellular damage irrespective of the attenuation of some cellular inflammatory
4 responses of the airways, pulmonary function, and symptom responses. Devlin et al.
5 (1997) found that several indicators of inflammation (e.g., PMN, IL-6, PGE2, fibronectin)
6 were attenuated after 5 days of exposure (i.e., values were not different from FA).
7 However, other markers (LDH, IL-8, total protein, epithelial cells) did not show
8 attenuation, suggesting that tissue damage probably continues to occur during repeated
9 exposure. Some cellular responses did not return to baseline levels for more than 10-20
10 days following O3 exposure. Christian et al. (1998) showed decreased numbers of
11 neutrophils and a decrease in IL-6 levels in healthy adults after 4 days of exposure versus
12 the single exposure to 0.2 ppm O3 for 4 h. Torres et al. (2000) also found both functional
13 and BALF cellular responses to O3 were abolished at 24 hours postexposure following
14 the fourth exposure day. However, levels of total protein, IL-6, IL-8, reduced glutathione
15 and ortho-tyrosine was still increased significantly. In addition, visual scores
16 (bronchoscopy) for bronchitis and erythema and the numbers of neutrophils in bronchial
17 mucosal biopsies were increased. Results indicate that, despite an attention of some
18 markers of inflammation in BALF and pulmonary function decrements, inflammation
19 within the airways persists following repeated exposure to O3. The continued presence of
20 cellular injury markers indicates a persistent effect that may not necessarily be recognized
21 due to the attenuation of spirometric and symptom responses.
22 A number of studies show that O3 exposures increase epithelial cell permeability through
23 direct (technetium-99m labeled diethylene triamine pentaacetic acid, 99mTc-DTPA,
24 clearance) and indirect (e.g., increased BALF albumin, protein) techniques. Kehrl et al.
25 (1987) showed increased 99mTc-DTPA clearance in healthy young adults at 75 minutes
26 postexposure to 0.4 ppm O3 for 2 hours. Foster and Stetkiewicz (1996) have shown that
27 increased 99mTc-DTPA clearance persists for at least 18-20 hours post-O3 exposure (130
28 minutes to average O3 concentration of 0.24 ppm), and the effect is greater at the lung
29 apices than at the base. Increased BALF protein, suggesting O3-induced changes in
30 epithelial permeability, have also been reported at 1 hour and 18 hours postexposure
31 (Devlin et al.. 1997; Balmes etal.. 1996). Meta-analysis of results from 21 publications
32 (Mudway and Kelly. 2004a). showed that increased BALF protein is associated with total
33 inhaled O3 dose (i.e., the product of O3 concentration, exposure duration, and VE).
34 It has been postulated that changes in permeability associated with acute inflammation
35 may provide increased access of inhaled antigens, particles, and other inhaled substances
36 deposited on lung surfaces to the smooth muscle, interstitial cells, and the blood. Hence,
37 increases in epithelial permeability following O3 exposure might lead to increases in
38 airway responsiveness to specific and nonspecific agents. Que et al. investigated this
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1 hypothesis in healthy young adults (83M, 55 F) exposed to 220 ppb O3 for 2.25 h
2 (alternating 15 min periods of rest and brisk treadmill walking). As has been observed by
3 others for FEVi responses, within individual changes in permeability were correlated
4 between sequential O3 exposures. This indicates differences in susceptibility to epithelial
5 damage from O3 exposure among individuals. Increases in epithelial permeability at 1
6 day post-O3 exposure were not correlated with FEVi responses immediately following
7 the O3 exposure nor with changes in airway responsiveness to methacholine in assessed 1
8 day post-O3 exposure. The authors concluded that changes in FEVi, permeability, and
9 airway responsiveness following O3 exposure were relatively constant over time in young
10 healthy adults; although, these responses appear to be mediated by differing physiologic
11 pathways.
6.2.3.2 Epidemiology
12 In the 2006 O3 AQCD, epidemiologic evidence of changes in pulmonary inflammation in
13 association with short-term ambient O3 exposure (30-min or 1-h max) was limited to
14 observations of increases in nasal lavage levels of inflammatory cell counts, eosinophilic
15 cationic protein, and myeloperoxidases (U.S. EPA. 2006b). As a result of the
16 development of less invasive methods to collect exhaled breath samples repeatedly from
17 individuals in the field, the number of studies assessing ambient O3-related changes in
18 lower airway inflammation and oxidative stress in recent years has increased
19 dramatically. Although most of the biomarkers examined in these studies were not
20 specific to the lung, most studies collected exhaled breath, exhaled breath condensate
21 (EEC), nasal lavage fluid, or induced sputum with the aim of monitoring inflammatory
22 responses in airways, as opposed to monitoring systemic responses in blood. These recent
23 studies form a larger base to establish coherence with findings from human experimental
24 and animal toxicological studies that have measured similar or related endpoints and
25 provide further biological plausibility for associations of ambient O3 exposure with
26 respiratory symptoms and lung function decrements. These biological markers also allow
27 the assessment of short-term O3-related acute respiratory effects in populations that are
28 less likely to experience respiratory symptoms, including healthy populations and groups
29 with increased outdoor exposures.
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Table 6-14 Mean and upper percentile ozone concentrations in studies
examining biological markers of pulmonary inflammation and
oxidative stress
Study
Qian et al.
(2009)
Khatrietal.
(2009)
Ferdinands etal.
(2008)
Adamkiewicz et
al. (2007)
Berhane et al.
(2011)
Delfino et al.
(201 Oa)
Liu et al. (2009a)
Sienra-Monge et
al. (2004)
Barraza-
Villarreal et al.
(2008)
Romieu et al.
(2008)
Location
Boston, MA; New York, NY;
Denver, CO; Philadelphia, PA;
San Francisco, CA; Madison, Wl
(SOCS)
Atlanta, GA
Atlanta, GA
Steubenville, OH
13 Southern California
Communities
Los Angeles, CA
Windsor, ON, Canada
Mexico City, Mexico
Mexico City, Mexico
Mexico City, Mexico
Years
1997-1999
All-year
2003, 2005,
2006
Warm season
2004
Warm season
2000
Cold season
September
2004- June
2005
2005-2007
All-year
2005
Cold season
1999-2000
All-year
2003-2005
All-year
2004
All-year
03
Averaging
Time
8-h max
8-h max
1-h max
24-h avg
1-havgb
8-h avg
(10:00 a.m. -
6:00 p.m.)
24-h avg
24-h avg
8-h max
8-h max
8-h max
Mean/Median
Concentration (ppb)
33.6
59a
61 (median)
15.3
19.8
NR
Warm season: 33.3
Cool season: 20.6
13.0
66.2
31.6
31.1
Upper Percentile
Concentrations (ppb)
75th: 44.4, Max
Max: 73a
75th: 67
75'": 20.2, Max:
75th: 27.5, Max:
NR
Max: 76.4
Max: 44.9
95": 26.5
Max: 142.5
Max (8-h max):
75'": 38.3
Max: 60.7
:91.5
32.2
61.6
86.3
Nickmilder et al. southern Belgium
(2007)
2002 1-h max NR
Warm season 8-h max NR
Max (across 6 camps):
24.5-112.7°
Max (across 6 camps):
18.9-81.1°
Chimenti et al. Sicily, Italy
NR 8-h avg Fall: 32.7 (pre-race), 35.1
All-year (7:00 a.m.- (race)0
3:00 p.m.) Winter: 37.0 (pre-race),
30.8 (race)0
Summer: 51.2 (pre-race),
46.1 (race)0
NR
Max = Maximum, NR = Not Reported.
'Individual-level exposure estimates were derived based on time spent in the vicinity of various 03 monitors.
bAverage 03 oncentration in the 1 h preceding eNO collection.
""Concentrations converted from ug/m3 to ppb using the conversion factor of 0.51 assuming standard temperature (25°C) and pressure (1 atm).
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Study
Population
O3Lag
Subgroup
Individuals with asthma
Liu et al. (2009) Children with asthma 0
Barraza-Villarreal et al.
(2008)
Children
Berhane et al. (2011) Children
0 Without asthma
With asthma
0-22 cum avg Without asthma
With asthma
Without allergy
With allergy
Qian et al. (2009)
Khatri et al. (2009)
Children and adults
with asthma
Adults with asthma 2
Older adults
Adamkiewicz et al. (2004) Olderadults
Delfino etal.(2010)
Olderadults
0-4 avg
Cool season
Warm season
-20 -10
10 20 30 40 50
Percent change in eNO per standardized increment in 03
(95% Cl)
Results are presented first for children with asthma followed by results for adults with asthma and older adults. Effect estimates
are from single-pollutant models and are standardized to a 30-ppb increase for 8-h max or 8-h avg ozone exposures and a 20-ppb
increase for 24-h avg ozone exposures.
Figure 6-10 Percent change in exhaled nitric oxide (eNO) per standardized
increment in ambient ozone exposure in studies of individuals with
and without asthma.
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Table 6-15 Additional characteristics and quantitative data for studies
represented in Figure 6-10
Study
Location/
Population
Os Lag Os Averaging Time
Subgroup
Standardized
Percent Change
(95% Cl)a
Studies in individuals with asthma
Liu et al. (2009a)
Barraza-Villarreal et al. (2008)
Berhaneetal. (2011)
Qianetal. (2009)
Khatri et al. (2009)
Windsor, ON, Canada
Children with asthma
Mexico City, Mexico
Children
12 Southern California communities
Children
6 U.S. communities
Children and adults with asthma
Atlanta, GA
Adults with asthma
0 24-h avg
0 8-h max
0-22 cum avg 8-h avg
(10:00a.m.-6:00p.m.)
0 8-h max
2 8-h max
Without asthma
With asthma
Without asthma
With asthma
Without allergy
With allergy
-17.0 (-30.3, -1.1)
13.5(11.2, 15.8)
6.2 (6.0, 6.5)
30.1 (10.6, 53.2)
26.0 (-1.4, 60.9)
25.5 (5.3, 49.6)
32.1 (12.0, 55.9)
-1.2 (-1.7, -0.64)
36.6(1.2,71.9)
Studies in older adults
Adamkiewicz et al. (2007)
Delfinoetal. (2010a)
Steubenville, Ohio
Older adults
Los Angeles, CA
Older adults
0 24-h avg
0-4 avg 24-h avg
Cool season
Warm season
-5.7 (-25.9, 14.5)
23.6 (7.3, 39.9)
-0.58 (-13.4, 12.3)
'Effect estimates are standardized to a 30-ppb increase for 8-h max or 8-h avg 03 and a 20-ppb increase for 24-h avg 03.
Table 6-16 Associations between short-term ambient ozone exposure and
biological markers of pulmonary inflammation and oxidative stress
Study
Liu et al. (2009a)
Romieu et al. (2008)
Barraza-Villarreal etal.
(2008)
Sienra-Monge et al.
(2004)
Khatri etal. (2009)
Ferdinands etal.
(2008)
Location/
Population
Windsor, ON, Canada
Children with asthma
Mexico City, Mexico
Children with asthma
Mexico City, Mexico
Children
Mexico City, Mexico
Children with asthma
Atlanta, GA
Adults with asthma
Atlanta, GA
Children exercising
outdoors
Os Os Averaging
Lag Time
0 24-h avg
0 8-h max
0 8-h max
0-2 avg 8-h max
2 8-h max
0 1-hmax
Biological Marker
EEC 8-isoprostane (%
change)
EEC TEARS (%
change)
EEC MDA"
Nasal lavage IL-8
(pg/ml)
Nasal lavage IL-8"
Nasal lavage IL-6b
Nasal lavage Uric acidb
Nasal lavage GSxb
Blood eosinophils (%
change)
EBCpH
Subgroup
Without asthma
With asthma
Without asthma
With asthma
Placebo
Antioxidant
Placebo
Antioxidant
Placebo
Antioxidant
Placebo
Antioxidant
Effect Estimate
(95% Cl)a
10.2 (-9.2, 33.5)
7.2 (-18.3, 40.7)
1 .3 (1 .0, 1 .7)
1 .6 (1 .4, 1 .8)
1 .6 (1 .4, 1 .9)
-0.10 (-0.27, 0.08)°
-0.10 (-0.20, 0.01)°
1.4 1.0,2.0)
1 .0 0.70, 1 .5)
1.5 1.2,2.0)
1 .0 (0.76, 1 .4)
0.88(0.70,1.1)
1.1 (0.84,1.5)
0.90 (0.82, 0.99)
0.91 (0.83, 0.98)
2.4 (0.62, 4.2)
2.5 (-0.20, 5.1)c
EEC = exhaled breath condensate, TEARS = thiobarbituric acid reactive substances, MDA = malondialdehyde, IL-8 = interleukin 8, IL-6 =
interleukin 6, Antioxidant=group supplemented with vitamins C and E, GSx = glutathione.
'Effect estimates are standardized to a 40-, 30- and 20-ppb increase for 1-h max, 8-h max, and 24-h avg 03, respectively
bModels analyzed log-transformed biological markers. Therefore, effect estimates represent the ratio of the geometric means of biological
markers for a standardized increase in 03 exposure. An estimate less than 1 indicates a decrease in pulmonary inflammation or oxidative stress for
an increase in 03 exposure, and an estimate greater than 1 indicates an increase in pulmonary inflammation or oxidative stress for an increase in
03 exposure.
°Negative and positive effect estimates indicate increases and decreases in pulmonary inflammation, respectively.
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1 Despite the strengths of biomarker studies, it is important to note that research in this
2 field continues to develop, and several uncertainties are recognized that may limit the
3 interpretations of associations between ambient O3 exposure and changes in biomarker
4 levels. Current areas of development include examination of the clinical relevance of the
5 observed magnitudes of changes in biological markers of pulmonary inflammation
6 (Murugan et al.. 2009; Duramad et al.. 2007). characterization of the time course of
7 changes between biomarker levels and other endpoints of respiratory morbidity,
8 development of standardized methodologies for collection, improvement of the
9 sensitivity and specificity of assay methods, and characterization of subject factors (e.g.,
10 asthma severity and recent medication use) that contribute to inter-individual variability.
11 These sources of uncertainty may contribute to differences in findings among studies.
12 In recent epidemiologic studies, the biomarker most frequently measured was exhaled
13 nitric oxide (eNO), likely related to its ease of collection in the field and automated
14 measurement. Other biological media analyzed included EEC, induced sputum, and nasal
15 lavage fluid, all of which are hypothesized to contain aerosolized particles and/or cells
16 from fluid lining the lower and upper airways (Balbi et al., 2007; Howarth et al., 2005;
17 Hunt. 2002). These fluids contain cytokines, cells, and markers of oxidative stress that
18 mediate inflammatory responses. In particular, several of the cytokines, cells, and
19 markers of oxidative stress examined in epidemiologic studies also have been examined
20 in controlled human exposure and toxicological studies.Table 6-14 presents the
21 characteristics and ambient O3 concentration data from recent studies assessing
22 associations between O3 exposure and biological markers of pulmonary inflammation and
23 oxidative stress. Many recent studies reported positive associations between short-term
24 ambient O3 exposure and increases in pulmonary inflammation and oxidative stress, in
25 particular, studies of children with asthma conducted in Mexico City (Figure 6-10 and
26 Tables 6-15 and 6-16).
Populations with Asthma
Exhaled Nitric Oxide
27 Nitric oxide or eNO has not been examined in controlled human exposure or
28 toxicological studies of O3 exposure. However, several lines of evidence support its
29 analysis as an indicator of pulmonary inflammation in epidemiologic studies. Inducible
30 nitric oxide synthase can be activated by pro-inflammatory cytokines, and NO can be
31 produced by cells such as neutrophils, eosinophils, and epithelial cells in the lung during
32 an inflammatory response (Barnes and Liew. 1995). Additional support is provided by
33 observations of higher eNO in individuals with asthma, and increases in the levels during
34 acute exacerbations (Jones et al.. 2001; Kharitonov and Barnes. 2000).
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1 As indicated in Figure 6-10 and Table 6-15, several studies found that short-term ambient
2 O3 exposure (8-h max or avg) was associated with increases in eNO in children with
3 asthma. Liu et al. (2009a) (described in Section 6.2.1.2) reported O3-associated decreases
4 in eNO; however, this study was restricted to winter months. In this study, results for
5 EEC levels of TEARS and 8-isoprostane as well as lung function did not provide strong
6 evidence of O3 effects on airway oxidative stress.
7 The two studies that compared children with and without asthma did not find larger O3-
8 associated increases in eNO in children with asthma (Figure 6-10 and Table 6-15).
9 Among children in Southern California, Berhane et al. (2011) examined a 0-22 day
10 cumulative average of 8-h avg (10:00 a.m.-6:00 p.m.) O3 and estimated similar effects for
11 children with and without asthma and children with and without allergy. Among children
12 in Mexico City, Barraza-Villarreal et al. (2008) examined lag 0 of 8-h max O3 and
13 estimated larger effects for children without asthma.
14 In the two studies that included adults with asthma, ambient O3 exposure was associated
15 with both decreases and increases in eNO (Khatri et al.. 2009; Qian et al.. 2009). In the
16 multicity salmeterol ((3-2 adrenergic agonist) trial (Boston, MA; New York, NY; Denver,
17 CO; Philadelphia, PA; San Francisco, CA; and Madison, WI), eNO was collected every
18 2-4 weeks over a 16-week period from 119 subjects with persistent asthma, 87% of
19 whom were 20-65 years of age (Qian et al.. 2009). Among all subjects, lag 0 of 8-h max
20 O3 was associated with a decrease in eNO as were exposures lagged 1 to 3 days and
21 averaged over 5 days. Results did not vary among the salmeterol, CS, and placebo
22 groups, indicating that the counterintuitive findings for O3 were not simply due to the
23 reduction of inflammatory responses by medication use. The authors suggested that at
24 higher O3 exposures, O3 may rapidly react with NO in airways to form reactive nitrogen
25 species such as peroxynitrite. However, in the other study of adults with asthma, ambient
26 concentrations of 8-h max O3 were higher, and a positive association was found with
27 eNO (Khatri et al.. 2009). In this study conducted during a summer season in Atlanta,
28 GA, a 30-ppb increase in lag 2 of 8-h max O3 was associated with a 36.6% increase in
29 eNO (95% CI: 1.2, 71.9). These findings should be interpreted with caution as they were
30 based on a single eNO measurement per subject and were not adjusted for any
31 meteorological factors.
Other biological markers of pulmonary inflammation and oxidative stress
32 As indicated in Table 6-16, studies have found associations between short-term ambient
33 O3 exposure and changes in the levels of proinflammatory cytokines and cells, indicators
34 of oxidative stress, and antioxidants. Importantly, any particular endpoint was examined
35 only in one to two studies, and the evidence in individuals with asthma is derived
36 primarily from studies conducted in Mexico City (Romieu et al.. 2009; Barraza-Villarreal
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1 et al.. 2008; Romieu et al., 2008; Sienra-Monge et al., 2004). Despite the limited
2 evidence, the epidemiologic observations are well-supported by controlled human
3 exposure and toxicological studies that have measured analogous endpoints.
4 Several of the modes of action of O3 are mediated by secondary oxidation products
5 produced in the airways by O3 (Section 5.3.3). Reactive oxygen species (ROS) are
6 involved in the regulation of inflammation via regulation of the expression of cytokines
7 and activity of inflammatory cells in airways (Heidenfelder et al.. 2009). In controlled
8 human exposure and toxicological studies, prostaglandins have been frequently measured
9 to indicate O3-induced increases in oxidative stress (Sections 5.3.3 and 6.2.3.1).
10 Prostaglandins are produced by the peroxidation of arachidonic acid in cell membranes
11 (Morrow et al.. 1990). Romieu et al. (2008) analyzed biweekly samples of EEC
12 malondialdehyde (MDA), a thiobarbituric acid reactive substance, which like
13 prostaglandins, is derived from oxidative degradation of lipids (Janero. 1990). The ratio
14 of the geometric means of MDA was 1.3 (1.0, 1.7) per a 30-ppb increase in lag 0 of 8-h
15 max O3. Similar results were reported for lags 1 and 0-1 average exposures. An important
16 limitation of the study was that 25% of EEC samples had nondetectable levels of MDA.
17 Thus, the random assignment of concentrations between 0 and 4.1 nmol may have
18 contributed random measurement error to the estimated O3 effects. Because MDA
19 represents less than 1% of lipid peroxides and is present at low concentrations, its
20 reliability as a marker of oxidative stress in vivo has been questioned. However, Romieu
21 et al. (2008) pointed to their observations of statistically significant associations of EEC
22 MDA levels with nasal lavage IL-8 levels to support its analysis as a biologically-
23 relevant indicator of pulmonary inflammation.
24 Uric acid and glutathione are ROS scavengers that are present in the airway ELF. While
25 the roles of these markers in the inflammatory cascade of asthma are not well
26 characterized, they are observed to be consumed in response to short-term O3 exposure in
27 controlled human exposure and animal studies (Section 5.3.3). Results from an
28 epidemiologic study also indicate that ambient O3 exposure may stimulate an antioxidant
29 response. In a panel study with three measurements of nasal lavage at 3-week intervals,
30 Sienra-Monge et al. (2004) found O3-associated decreases in nasal lavage levels of uric
31 acid and glutathione in children with asthma not supplemented with antioxidant vitamins
32 (Table 6-16). The magnitude of association was similar for O3 exposures lagged 2 or 3
33 days and averaged over 3 days.
34 Both controlled human exposure and toxicological studies find O3-induced increases in
35 the cytokines IL-6 and IL-8 (Sections 5.3.3, 6.2.3.1, and 6.3.3.3), which are involved in
36 initiating an influx of neutrophils, a hallmark of inflammation induced by short-term O3
37 exposure. Recent epidemiologic studies produced similar findings. Barraza-Villarreal
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1 et al. (2008) observed that a 30-ppb increase in lag 0 of 8-h max O3 was associated with a
2 1.61 pg/ml increase (95% CI: 1.4, 1.8) in IL-8. In another study of children with asthma
3 in Mexico City, Sienra-Monge et al. (2004) found that lags 2, 3, and 0-2 avg of 8-h max
4 O3 were associated with increases in nasal lavage levels of IL-6 and IL-8 (placebo group),
5 with the largest effects estimated for lag 0-2 average exposure (Table 6-16).
6 Neutrophil influx has been a prominent characterisitc of O3-induced inflammation;
7 however, controlled human exposure studies also have found O3-induced increases in
8 eosinophils in adults with asthma (Section 6.2.3.1). Eosinophils are believed to be the
9 main effector cells that initiate and sustain inflammation in asthma and allergy (Schmekel
10 et al., 2001). Consistent with these findings, in a cross-sectional study of adults with
11 asthma in Atlanta, GA, a 30-ppb increase in lag 0 of 8-h max O3 was associated with a
12 2.4% increase (0.62, 4.2) in blood eosinophils (Khatri et al.. 2009). These results were
13 not adjusted for meteorological factors.
14 The pH of EEC also was analyzed as an indicator of pulmonary inflammation. EEC pH is
15 thought to reflect the proton-buffering capacity of ammonium in airways. It has been
16 widely used in the clinical assessment of asthma, is consistently lower in subjects with
17 asthma, decreases upon acute asthma exacerbation (on the order of 2 units), and is
18 negatively correlated with airway levels of proinflammatory cytokines (Carpagnano et
19 al.. 2005: Kostikas et al.. 2002: Hunt et al.. 2000). In addition to finding O3-associated
20 increases in eNO and nasal lavage IL-8, Barraza-Villarreal et al. (2008) found small O3-
21 associated decreases in EEC pH (Table 6-16).
22 The prominent influences of ROS and antioxidants in mediating the effects of O3 provide
23 biological plausibility for the effect modification by antioxidant supplementation. The
24 modulation of O3-associated lung function by antioxidant capacity has been described in
25 controlled human exposure and epidemiologic studies (Sections 6.2.1.1 and 6.2.1.2).
26 Epidemiologic studies also found that higher levels of dietary or supplemented
27 antioxidants attenuated inflammation and oxidative stress. Sienra-Monge et al. (2004)
28 conducted a 12 week-trial with daily vitamin C and E supplements. In the antioxidant
29 group, the ratios of the geometric means of nasal lavage IL-6 and IL-8 per 30-ppb
30 increases in lag 0-2 avg of 8-h max O3 were 1.0, reflecting no increases with increases in
31 O3 exposure (Table 6-16). Effect modification by antioxidant supplementation was not
32 consistent for uric acid and glutathione (Table 6-16). Ozone was associated with
33 increases in uric acid in the antioxidant group and decreases in the placebo group across
34 O3 lags of exposure. Associations with glutathione were similar in both groups.
3 5 Therefore, the results do not clearly delineate the interactions among inhaled O3,
36 endogenous antioxidants, and dietary supplementations of antioxidants. In another cohort
37 of children with asthma in Mexico City, a diet high in fruits and vegetables was found to
Draft - Do Not Cite or Quote 6-73 September 2011
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1 protect against O3-related increases in nasal lavage IL-8 (Romieu et al., 2009). At high
2 ambient O3 levels (> 38 ppb, 8-h max), a 1-unit increase in FVI was associated with a
3 0.219 pg/ml decrease (95% CI: -0.38, -0.05) in IL-8. The protective effect was
4 diminished by about 49% at O3 levels of 25 ppb or lower. Results from these two studies
5 indicate that augmenting the circulating levels of antioxidants, through diet or vitamin
6 supplements, may reduce nasal inflammation associated with high ambient O3 exposures.
Clinical significance of ozone-associated changes in pulmonary
inflammation and oxidative stress in children with asthma
7 While the results of epidemiologic studies in children with asthma were consistent with
8 the known modes of action of O3 in consuming antioxidants and inducing oxidative stress
9 and pulmonary inflammation (Section 5.3.3), the clinical significance of these changes
10 has not been well-characterized. The levels of several of the biological markers such as
11 eNO, EEC pH, and MDA have been shown to differ between subjects with and without
12 asthma and change acutely during an acute asthma exacerbation (Corradi et al., 2003;
13 Hunt et al. 2000): however, the magnitudes of change for these conditions are not well-
14 defined. Several studies conducted in individuals with asthma found large O3-associated
15 increases in eNO; effect estimates ranged between a 6 and 36% increase per standardized
16 increment in ambient O3 concentration1 (Figure 6-10 and Table 6-15). Standardized
17 increments in ambient O3 exposure were associated with smaller (1-2%) increases in
18 interleukins or indicators of oxidative stress (Khatri et al., 2009; Barraza-Villarreal et al..
19 2008) (Romieu et al.. 2008; Sienra-Monge et al.. 2004).
20 Some studies permitted the evaluation of the potential clinical relevance of these changes
21 in eNO through the concurrent assessment of respiratory symptoms. Among children
22 with asthma in Mexico City, O3 exposure was associated with increases in eNO and nasal
23 lavage IL-8 and concurrently assessed cough but not wheeze (Barraza-Villarreal et al.,
24 2008). Among adults with asthma in Atlanta, O3 was associated with increases in eNO,
25 blood eosinophils, and a decrease in quality of life score, which incorporates indices for
26 symptoms, mood, and activity limitations (Khatri et al.. 2009). These findings suggest
27 that the more subtle O3-associated increases in biological markers of airway
28 inflammation may be sufficient to result in respiratory symptoms or activity limitations.
Children without Asthma
29 Recent studies found that short-term O3 exposure (8-h max or avg) was associated with
30 indicators of airway inflammation in children without asthma (Berhane et al., 2011;
31 Barraza-Villarreal et al.. 2008) (Figure 6-10 and Tables 6-15 and 6-16). In the panel
1 Effect estimates were standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max, and 24-h avg O3.
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1 study of children in Mexico City, O3 exposure was associated with a larger increase in
2 eNO in the children without asthma than with asthma (13.5% versus 6.2% increase per
3 30-ppb increase in lag 0 of 8-h max O3). Ozone was associated with similar magnitudes
4 of changes in IL-8 and EEC pH in children with and without asthma. A distinguishing
5 feature of this study was that most of the children without asthma were atopic (72%) as
6 indicated by positive skin prick tests, which may have contributed to the similar effects of
7 O3 exposure observed in children with and without asthma.
8 However, the Southern California Children's Health Study estimated similar effects for
9 8-h avg (10:00 a.m.-6:00 p.m.) ambient O3 exposure on eNO in children with and without
10 respiratory allergy (Berhane et al., 2011). Results from this large study (n = 2240
11 children) provided evidence that ambient O3 exposure increases airway inflammation in
12 healthy children. In comparison with other studies, this analysis from the Children's
13 Health Study provided detailed information on differences in association among various
14 lags of 8-h avg (10:00 a.m.-6:00 p.m.) O3 exposure. Consistent with other studies
15 examining pulmonary inflammation and oxidative stress, Berhane et al. (2011) found that
16 relatively shorter lags of exposure, including 1 to 5 days, were associated with increases
17 in eNO. However, in an examination of several types of lag-based models, including
18 unconstrained lag models, polynomial distributed lag models, spline-based distributed lag
19 models, and cumulative lag models, investigators found that a 23-day cumulative lag
20 model best fit the data. Among the studies evaluated in the current assessment, Berhane
21 et al. (2011) was unique in evaluating and finding larger effects for cumulative average
22 O3 exposures over multiple weeks (e.g., 13-30 days). O3 exposures averaged over the
23 several hours preceding eNO collection were not significantly associated with eNO. The
24 mechanism for the effects of O3 peaking with a 23-day cumulative lag of exposure is not
25 known.
Populations with Increased Outdoor Exposures
26 In a limited number of available studies, ambient O3 exposure was not consistently
27 associated with pulmonary inflammation in populations with increased outdoor
28 exposures. Important limitations of these studies include small numbers of subjects and
29 repeated measurements. In a cross-sectional study of children at camps in south Belgium,
30 although O3 was not associated with lung function, an association was found for eNO
31 (Nickmilder et al.. 2007). Children at camps with lag 0 1-h max O3 concentrations above
32 85.2 ppb had greater increases in intraday eNO compared with children at camps with O3
33 concentrations below 51 ppb. A benchmark dose analysis indicated that the threshold for
34 an O3-induced increase of 4.3 ppb eNO (their definition of increased pulmonary
35 inflammation) was 68.6 ppb for 1-h max O3 and 56.3 ppb for 8-hr max O3. While these
36 results provide additional evidence for O3-associated increases in airway inflammation in
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1 healthy children, they should be interpreted with caution since they were not adjusted for
2 any potential confounding factors.
3 Recent studies examined associations of O3 exposure with biological markers of airway
4 inflammation in populations exercising outdoors. In a panel study of 16 adolescent long-
5 distance runners in Atlanta, GA, lags 0, 1, and 2 of 1-h max O3 were associated with
6 increases in EEC pH, indicating O3-associated decreases in pulmonary inflammation
7 (Ferdinands et al.. 2008). Among 9 adult male runners in Sicily, Italy examined 3 days
8 before and 20 hours after 3 races in fall, winter, and summer, weekly average O3
9 concentrations (8-h avg, 7:00 a.m.-3:00 p.m.) were positively correlated with apoptosis of
10 airway cells (Spearman's r = 0.76, p < 0.0005) and bronchial epithelial cell differential
11 counts (Spearman's r = 0.467, p < 0.05) but not with neutrophil or macrophage cell
12 counts or levels of the proinflammatory cytokines TNF-a and IL-8 (Chimenti et al..
13 2009). Although this study provides evidence for some new endpoints, the implications
14 of the findings are limited since they were not based on a rigorous statistical analysis.
Older Adults
15 Two panel studies examining O3-associated changes in eNO in older adults produced
16 contrasting findings (Figure 6-10 and Table 6-15). Both studies were similar in that
17 outdoor O3 was monitored by investigators in the vicinity of subjects' residences, and
18 cool season-specific results were presented. However, several differences were
19 noteworthy, including geographic location, inclusion of healthy subjects, and lags of O3
20 exposure examined. Delfino et al. (2010a) followed 60 elderly subjects with coronary
21 artery disease in the Los Angeles, CA area for two 6-week periods, one in the warm
22 season and one in the cool season, although the exact months were not specified.
23 Multiday averages of O3 (3- to 9-day) were associated with increases in eNO, with effect
24 estimates increasing with increasing number of averaging days. In contrast with most
25 other studies, a strong positive effect was estimated for the cooler season (4.06 ppb [95%
26 CI: 1.25, 6.87]) increase in eNO per 20-ppb increase in lag 0-4 of 24-h avg O3), whereas
27 no association was observed for the warm season (-0.01 ppb change in eNO [95% CI: -
28 2.31, 2.11]). Despite these unusual findings for the cool season, they were similar to
29 findings from another study of Los Angeles area adults with asthma that found O3 effects
30 (i.e., decrease in indoor activity) during the fall season (Eiswerth et al.. 2005).
31 Adamkiewicz et al. (2004) did not find a positive association between O3 exposure and
32 eNO in a group of older adults (ages 54-91 years) comprising healthy subjects and those
33 with asthma or COPD. The study was conducted in Steubenville, OH between September
34 and December, and as was observed in most other studies conducted during winter
35 months, O3 (concurrent 1 hour and 24 hours preceding eNO collection) was associated
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1 with decreases in eNO, indicating a decrease in pulmonary inflammation (Figure 6-10
2 and Table 6-16).
Confounding in Epidemiologic Studies of Pulmonary Inflammation and
Oxidative Stress
3 Except where noted in the preceding text, most epidemiologic studies of pulmonary
4 inflammation and oxidative stress accounted for the potential for confounding by
5 meteorological factors. Ambient O3 exposure was associated with pulmonary
6 inflammation or oxidative stress in models that adjusted for temperature and/or humidity
7 (Delfinoetal..2010a: Barraza-Villarreal et al.. 2008; Romieu et al.. 2008). Most studies
8 conducted over multiple seasons adjusted for season or time trend. Sienra-Monge et al.
9 (2004) and Berhane et al. (2011) did not adjust for temperature in their final results after
10 finding that the inclusion of temperature did not change results.
11 Although information is limited to a small number of studies conducted in Mexico City,
12 the evidence does not indicate the confounding of O3 associations by PM2 5 or PM10
13 exposure. In these studies, which analyzed 8-h averages for both O3 and PM and reported
14 moderate correlations between pollutants (r=0.46-0.54), robust associations were found
15 for O3 (Barraza-Villarreal et al.. 2008: Romieu et al.. 2008: Sienra-Monge et al.. 2004).
16 Only Romieu et al. (2008) provided quantitative results. Lag 0 of 8-h max O3 was
17 associated with the same magnitude of increase in MDA with and without lag 0 of 8-h
18 max PM25 in the model (ratio of geometric means per 30-ppb increase: 1.3 [95% CI: 1.0,
19 1-7]). In the copollutant model, the effect estimate for PM2s was cut in half.
Summary of Epidemiologic Studies of Pulmonary Inflammation and
Oxidative Stress
20 Many recent epidemiologic studies reported positive associations between short-term
21 ambient O3 exposure and increases in pulmonary inflammation and oxidative stress,
22 particularly, studies of children with asthma in Mexico City. By also finding that O3-
23 associated increases in pulmonary inflammation were attenuated with higher antioxidant
24 intake, these studies, as a whole, provided evidence that inhaled O3 may be an important
25 source of ROS in airways and/or may increase airway inflammation via oxidative stress-
26 mediated mechanisms. Studies also indicated that ambient O3 exposure may increase
27 airway inflammation in healthy children (Berhane et al.. 2011: Nickmilder et al.. 2007).
28 The limited available evidence in subjects exercising outdoors and older adults was
29 inconclusive. Temperature and humidity were not found to confound O3 associations, and
30 in the few studies that evaluated copollutant models, O3 effect estimates were robust to
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1 inclusion of PM7 s or PMin (Barraza-Villarreal et al., 2008; Romieu et al., 2008; Sienra-
2 Mongeetal.. 2004).
3 Most studies examined associations with daily 8-h max or daytime 8-h avg O3 exposures,
4 although associations were observed for 1-h max (Nickmilder et al., 2007) and 24-h avg
5 O3 exposures (Delfino et al.. 2010a). Collectively, studies examined associations with
6 single-day O3 exposures lagged from 0 to 5 days, and exposures averaged over 2 to 9
7 days. Lag 0 of 8-h max O3 exposure was most frequently examined and consistently
8 associated with increased airway inflammation and oxidative stress. However, in the few
9 studies that examined multiple lags of exposure, multiday cumulative O3 exposures,
10 primarily based on 8-h max or 8-h avg, were associated with greater increases in airway
11 inflammation and oxidative stress (Berhane et al.. 2011; Delfino et al.. 2010a: Sienra-
12 Monge etal., 2004). These findings for longer lags of exposure are supported by
13 controlled human exposure studies that similarly have found that indicators of airway
14 inflammation remain elevated following exposures to O3 repeated over multiple days
15 (Section 6.2.3.1).
16 Several epidemiologic studies simultaneously examined associations of ambient O3
17 exposure with biological markers of airway inflammation and oxidative stress, lung
18 function, and respiratory symptoms. In most cases, the results differed between the
19 various biomarkers and lung function. Whether evaluated at the same or different lags of
20 O3 exposure, associations generally were stronger for biological markers of airway
21 inflammation than for lung function (Barraza-Villarreal et al.. 2008; Nickmilder et al..
22 2007). Controlled human exposure studies also have demonstrated a lack of correlation
23 between inflammatory and spirometric responses induced by O3 exposure. Studies have
24 suggested that O3-related respiratory morbidity may occur via multiple mechanisms with
25 varying time courses of action, and the examination of a limited number of O3 exposure
26 lags in these aforementioned studies may explain some of the inconsistencies in
27 associations of O3 with different respiratory health endpoints.
28 The clinical significance of changes in biological markers of airway inflammation and
29 oxidative stress are not well-characterized. However, the simultaneous examination of
30 associations of O3 with respiratory symptoms has permitted the assessment of the clinical
31 significance of the changes observed in biomarkers. In subjects with asthma, ambient O3
32 exposure was associated with increases in eNO and IL-6 that were accompanied by a
33 concomitant increase in cough (Barraza-Villarreal et al., 2008) and increases in eNO and
34 blood eosinophils that were accompanied by a decrease in quality of life score (Khatri et
35 al., 2009). These findings support clinically-important increases in O3-associated airway
36 inflammation in individuals with asthma. Similar data are limited to assess the clinical
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1 significance of changes in other biological markers of airway inflammation and oxidative
2 stress and in other populations.
6.2.3.3 Toxicology
3 The 2006 O3 AQCD states that the "extensive human clinical and animal toxicological
4 evidence, together with the limited available epidemiologic evidence, is clearly indicative
5 of a causal role for O3 in inflammatory responses in the airways" (U.S. EPA. 2006b).
6 Airway ciliated epithelial cells and Type 1 cells are the most O3-sensitive cells and are
7 initial targets of O3. These cells are damaged by O3 and produce a number of
8 proinflammatory mediators (e.g., interleukins [IL-6, IL-8], PGE2) capable of initiating a
9 cascade of events leading to PMN influx into the lung, activation of alveolar
10 macrophages, inflammation, and increased permeability across the epithelial barrier. One
11 critical aspect of inflammation is the potential for metaplasia and alterations in
12 pulmonary morphology. Studies have observed increased thickness of the alveolar septa,
13 presumably due to increased cellularity after acute exposure to O3. Epithelial hyperplasia
14 starts early in exposure and increases in magnitude for several weeks, after which it
15 plateaus until exposure ceases. When exposure persists for a month and longer, excess
16 collagen and interstitial fibrosis are observed. This response, discussed in Chapter 7,
17 continues to increase in magnitude throughout exposure and can even continue to
18 increase after exposure ends (Last et al.. 1984). Previously published toxicological
19 studies of the ability of O3 to cause inflammation, injury, and morphological changes are
20 described in Table 6-5 on p. 6-25 and Tables 6-10 and 6-11 beginning on p. 6-61 of the
21 1996 O3 AQCD, and Tables AX5-8 and AX5-9, beginning on p. AX5-17 of the 2006 O3
22 AQCD. Numerous recent in vitro and in vivo studies add to this very large body of
23 evidence for O3-induced inflammation and injury, and provide new information regarding
24 the underlying mechanisms (Bauer et al.. 2011; Aibo et al.. 2010; Farraj et al.. 2010;
25 Garantziotis et al.. 2010; Hicks etal.. 2010b: Castagna et al.. 2009; Damera et al.. 2009;
26 Oslund et al.. 2009: Vancza et al.. 2009: Vovnow et al.. 2009: Fakhrzadeh et al.. 2008:
27 Han et al.. 2008: Inoue et al.. 2008: Oslund et al.. 2008: Carey et al.. 2007: Cho et al..
28 2007: Dahl et al.. 2007: Johnston et al.. 2007: Kooter et al.. 2007: Wagner etal.. 2007:
29 Wang et al.. 2007: Yoon et al.. 2007: Huffman et al.. 2006: Johnston et al.. 2006: Kenyon
30 et al.. 2006: Manzer et al.. 2006: Plopper et al.. 2006: Jang et al.. 2005: Janic et al.. 2005:
31 Johnston et al.. 2005a: Johnston et al.. 2005b: Oyarzun etal.. 2005: Servais etal.. 2005:
32 Frush et al.. In Press).
33 A number of species, including dogs, rabbits, guinea pigs, rats, and mice have been used
34 as models to study the pulmonary effects of O3, but the similarity of non-human primates
35 to humans makes them an attractive model in which to study the pulmonary response to
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1 O3. As reviewed in the 1996 and 2006 O3 AQCDs, several pulmonary effects, including
2 inflammation, changes in morphometry, and airway hyperresponsiveness, have been
3 observed in macaque and rhesus monkeys after acute exposure to O3 (Table 6-17 presents
4 a highlight of these studies). Increases in inflammatory cells were observed after a single
5 8-hr exposure of adult rhesus monkeys to 1 ppm O3 (Hyde et al., 1992). Inflammation
6 was linked to morphometric changes, such as increases in necrotic cells, smooth muscle,
7 fibroblasts, and nonciliated bronchiolar cells, which were observed in the trachea,
8 bronchi, or respiratory bronchioles. Effects have also been observed after short-term
9 repeated exposure to O3 at concentrations that are more relevant to ambient O3 levels.
10 Morphometry changes in the lung, nose, and vocal cords were observed after exposure to
11 0.15 ppm O3 for 8-h/day for 6 days (Harkema et al.. 1993; Dimitriadis. 1992; Harkema et
12 al.. 1987a). Since 2006, however, only one study has been published regarding acute
13 exposure of non-human primates to O3 (a number of recent chronic studies in non-human
14 primates are described in Chapter 7). In this study, a single 6-h exposure of adult male
15 cynomolgus monkeys to 1 ppm O3 induced significant increases in inflammatory and
16 injury markers, including BAL neutrophils, total protein, alkaline phosphatase, IL-6, IL-
17 8, and G-CSF (Hicks et al., 2010b). Gene expression analysis confirmed the increases in
18 the pro-inflammatory cytokine IL-8, which had been previously described in O3 exposed
19 rhesus monkeys (Chang et al.. 1998). The anti-inflammatory cytokine IL-10 was also
20 elevated, but the fold changes in IL-10 and G-CSF were relatively low and highly
21 variable. The single exposure also caused necrosis and sloughing of the epithelial lining
22 of the most distal portions of the terminal bronchioles and the respiratory bronchioles.
23 Bronchiolitis, alveolitis, parenchymal and centriacinar inflammation were also observed.
24 A second exposure protocol (two exposures with a 2-week inter-exposure interval)
25 resulted in similar inflammatory responses, with the exception of total protein and
26 alkaline phosphatase levels which were attenuated, indicating that attenuation of some
27 but not all lavage parameters occurred upon repeated exposure of non-human primates to
28 O3 (Hicks et al.. 201 Ob). This variability in adaptation is similar to the findings of earlier
29 reports in rodents (Wiester et al.. 1996b) and non-human primates (Tyler etal.. 1988).
30 Table 6-17 describes morphometric studies conducted in non-human primates exposed
31 toO3.
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Table 6-17 Morphometric observations in non-human primates after acute O3
exposure
Reference
Harkema et al.
(1993)
Harkema et al.
(1987a; 1987b)
Dimitriadis (1992)
Leonard et al. (1991)
Chang et al. (1998)
Hyde et al. (1992)
Hicks et al. (201 Oa)
Os concentration
0.15
0.15
0.3
0.25
0.96
0.96
1.0
Exposure
duration
8 h/day for 6
days
8 h/day for 6
days
8 h/day for 7
days
8h
8h
6h
Species, Sex, Age
Macaca radiata
Macaca radiata, M, F
2-6 years old
Macaca radiata
Rhesus, M
Rhesus, M
2-8.5 years old
Cynomolgus, M
5-7 kg
Observation
Several fold increase in thickness of surface
epithelium in respiratory bronchioles
Ciliated cell necrosis, shortened cilia, and
increased mucous cells in the respiratory
epithelium of nose after 0.1 5 ppm; changes in
nonciliated cells, intraepithelial leukocytes, and
mucous cells in the transitional epithelium
The 03 exposure level is not clear - the abstract
states 0.64 ppm, but the text mentions only 0.25
ppm. Morphometric changes in vocal cord mucosa:
disruption and hyperplasia of stratified squamous
epithelium; epithelial and connective tissue
thickness increased
Increase in IL-8 in airway epithelium correlated with
PMN influx
Increased PMNs; morphometric changes in
trachea, conducting airways, respiratory
bronchioles
Increase in PMNs and IL-8 in lavage fluid
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Confirmation of pulmonary changes observed in non-human primates, at near ambient
(^concentrations, has been done in a large number of studies in guinea pigs and rodents
(see 1996 and 2006 O3 AQCDs) (U.S. EPA. 2006b. 1996a). Mechanistic studies
completed more recently have extended these findings. Exposure of adult BALB/c mice
to 0.1 ppm O3 for 4 hours increased BAL levels of keratinocyte chemoattractant (KC; IL-
8 homologue) (~ sixfold), IL-6 (~12-fold), and TNF-a (~ twofold) (Damera et al.. 2010).
Additionally, O3 increased BAL neutrophils by 21% without changes in other cell types.
A trend of increased neutrophils with increased O3 concentration (0.12-2 ppm) was
observed in BALB/c mice exposed for 3 hours (Jang et al.. 2005). Although alterations in
the epithelium of the airways were not evident in 129J mice after 4 hours of exposure to
0.2 ppm O3 (Plopper et al.. 2006). detachment of the bronchiolar epithelium was
observed in SD rats after 5 days or 60 days of exposure to 0.25 ppm O3 (Oyarzun et al..
2005). Subacute (65 hours) exposure to 0.3 ppm O3 induced pulmonary inflammation,
cytokine induction, and enhanced vascular permeability in wild type mice of a mixed
background (129/Ola and C57BL/6) and these effects were exacerbated in
metallothionein I/II knockout mice (Inoue et al.. 2008). Three hours or 72 hours of
exposure to 0.3 ppm O3 resulted in similar levels of IL-6 expression in the lungs of
C57BL/6 mice (Johnston et al.. 2005b). along with increases in BAL protein, sTNFRl,
and sTNFR2. Increased neutrophils were observed only after the 72-h exposure, and
neither exposure resulted in detectable levels of IL-6 or KC protein. Levels of BAL
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1 protein, sTNFRl, and sTNFR2 were higher in the 72-h exposure group than in the 3-h
2 exposure group. In another study, the same subacute (72 hours) exposure protocol elicited
3 increases in BALF protein, IP-10, sTNFRl, macrophages, neutrophils, and IL-6, IL-la,
4 and IL-1(3 expression (Johnston et al.. 2007). Yoon et al. (2007) exposed C57BL/6J mice
5 continuously to 0.3 ppm O3 for 6, 24, 48, or 72 hours, and observed elevated levels of
6 KC, MIP-2, metalloproteinases, and inflammatory cells in the lungs at various time
7 points. A similar exposure protocol using C3FI/HeJ and C3FI/OuJ mice demonstrated
8 elevations in protein, PMNs, and KC, which were predominantly TLR 4 pathway
9 dependent based on their prominence in the TLR 4 sufficient C3FI/OuJ strain (Bauer et
10 al.. 2011). C3H/OuJ mice also had elevated levels of the heat-shock protein HSP70, and
11 further experiments in HSP70 deficient mice indicated a role for this particular pathway
12 in O3-related injury, discussed in more detail in Chapter 5.
13 As reviewed in the 2006 O3 AQCD, the time course for changes in BAL depends on the
14 parameters being studied. Similarly, after exposing adult C57BL mice to 0.5 ppm O3 for
15 3 hours, Han et al. (2008) observed early (5 hours postexposure) increases in BAL TNF-a
16 and IL-lp, which diminished by 24 hours postexposure. Total BAL protein was elevated
17 at 24 hours, but there were only minimal or negligible changes in LDH, total cells, or
18 PMNs. Ozone increased BAL mucin levels (with statistical significance by 24 hours
19 postexposure), and significantly elevated surfactant protein D at both time points. Prior
20 intratracheal (IT) exposure to multiwall carbon nanotubes enhanced most of these effects,
21 but the majority of responses to the combined exposure were not greater than those to
22 nanotubes alone. Ozone exposure did not induce markers of oxidative stress in lung
23 tissue, BAL, or serum. Consistent with this study, Aibo et al. (2010) did not detect
24 changes in BAL inflammatory cell numbers in the same mouse strain after a 6-h exposure
25 to 0.25 or 0.5 ppm. The majority of inflammatory cytokines (pulmonary or circulating)
26 were not significantly changed (as assessed 9 hours post O3 exposure).
27 Animal toxicology studies have also examined susceptibility factors and the findings
28 complement research in both controlled human exposure and epidemiologic studies. In a
29 study examining age, strain, and gender as factors for susceptibility to O3 in mice,
30 increased BAL neutrophils were observed in all 8 strains of neonates and adults but
31 statistical significance was found in only 4 strains of neonates and 2 strains of adults at
32 24 hours after exposure to 0.8 ppm O3 for 5 hours (Vancza et al., 2009). Lung injury, as
33 measured by BAL protein, was significantly increased in 5 and 8 strains of neonates and
34 adults, respectively. Interestingly, the observed age-dependent differences in response to
35 O3 occurred in only certain strains. For example, the fold-increase in neutrophils was
36 significantly higher, in neonates compared to adults, in the SJL and C3H/HeJ strains and
37 lower in BALB/c mice. Measurement of 18O determined that the observed strain- and
38 age-dependent differences were not due to absorbed O3 dose. Subanalysis of the adult
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1 mice demonstrated that gender also played a small, but statistically significant, role in the
2 effect of O3 on BAL neutrophils and protein. These findings suggest that the response to
3 O3, in mice, may consist of a complex interaction of age, gender, and genetic factors.
4 A study assessing NQO1 as a susceptibility factor was conducted by Voynow et al.
5 (2009). Specific effects of this gene on O3 responses are discussed in Chapter 8; only
6 ozone's effects in wild type C57BL/6 mice are described here. Exposure to 1 ppm for
7 3 hours increased BAL total cells, neutrophils, and KC; these responses were greatest at
8 24 hours postexposure. F2-isoprostane (8-isoprostane), a marker of oxidative stress, was
9 also elevated by O3, peaking at 48 hours postexposure.
10 Atopic asthma appears to be a risk factor for more severe O3 induced airway
11 inflammation in humans (Balmes et al., 1997; Scannell et al., 1996). and allergic animal
12 models are often used to investigate the effects of O3 on this susceptible population.
13 Farraj et al. (2010) exposed allergen-sensitized adult male BALB/c mice to 0.5 ppm O3
14 for 5 hours once per week for 4 weeks. Ovalbumin-sensitized mice exposed to O3 had
15 significantly increased BAL eosinophils by 85% and neutrophils by 103% relative to
16 OVA sensitized mice exposed to air, but these changes were not evident upon
17 histopathological evaluation of the lung, and no O3 induced lesions were evident in the
18 nasal passages. Ozone increased BAL levels of N-acetyl-glucosaminidase (NAG; a
19 marker of injury) and protein. DEP co-exposure (2.0 mg/m3, nose only) inhibited these
20 responses. These pro-inflammatory effects in an allergic mouse model have also been
21 observed in rats. Wagner et al. (2007) exposed the relatively O3-resistant Brown Norway
22 rat strain to 1 ppm O3 after sensitizing and challenging with OVA. Rats were exposed for
23 2 days, and airway inflammation was assessed one day later. Filtered air for controls
24 contained less than 0.02 ppm O3. Histopathology indicated O3 induced site-specific lung
25 lesions in the centriacinar regions, characterized by wall thickening partly due to
26 inflammatory cells influx. BAL neutrophils were elevated by O3 in allergic rats, and
27 modestly increased in non-allergic animals (not significant). A slight (but not significant)
28 increase in macrophages was observed, but eosinophil numbers were not affected by O3.
29 Soluble mediators of inflammation (Cys-LT, MCP-1, and IL-6) were elevated by O3 in
30 allergic animals but not non-allergic rats. Treatment with yT, which neutralizes oxidized
31 lipid radicals and protects lipids and proteins from nitrosative damage, did not alter the
32 morphologic character or severity of the centriacinar lesions caused by O3, nor did it
33 reduce neutrophil influx. It did, however, significantly reduce O3-induced soluble
34 inflammatory mediators in allergic rats. The effects of O3 in animal models of allergic
35 asthma are discussed in section 6.2.6.
36 In summary, a large number of toxicology studies have demonstrated that acute exposure
37 to O3 produces injury and inflammation in the mammalian lung, supporting the
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1 observations in controlled human exposure studies (Section 6.2.3.1). These acute
2 changes, both in inflammation and morphology, provide a modicum of evidence for long
3 term sequelae of exposure to O3. Related alterations resulting from long term exposure,
4 such as fibrotic changes, are discussed in Chapter 7.
Mechanisms of Injury
5 Since O3 has been well established as a causative agent of airway inflammation and
6 injury, which may contribute to functional changes observed in human subjects, the
7 majority of recent research has focused on the underlying mechanisms. A brief
8 description of some of the recent contributions to this area of research is provided here;
9 more detailed descriptions of the mechanisms behind O3-mediated injury and
10 inflammation can be found in the mode of action chapter (Chapter 5). There are several
11 signaling pathways responsive to changes in oxidation status, which tend to be influenced
12 at different levels in different host backgrounds. The molecular mechanisms of TNF
13 receptor-mediated lung injury induced by O3 and associated signaling pathways (NF-KB,
14 MAPK/AP-1) have been examined (Fakhrzadeh et al.. 2008; Cho et al.. 2007). along with
15 the changes in gene expression which characterize O3-induced stress and inflammation
16 (Wang et al.. 2007). Other contributors to injury and inflammation include the IL-1 and
17 neurokinin receptors (Oslund et al.. 2008; Johnston et al.. 2007). calcitonin gene-related
18 peptide receptor activation (Oslund et al.. 2009). CXCR2, a receptor for neutrophil
19 chemokines (Johnston et al.. 2005a). mindin, an extracellular matrix protein (Frush et al..
20 In Press), and NQO1 (Voynow et al.. 2009). an enzyme involved in oxidative stress.
21 Studies indicate a role for oxidative stress in mediating inflammation (Wagner et al..
22 2007; Jang et al.. 2005). Protective roles have been identified for nitric oxide synthase
23 (Kenyon et al.. 2006). metallothionein (Inoue et al.. 2008). matrix metalloproteinases
24 (Yoon et al.. 2007). Clara cell secretory protein (Plopper et al.. 2006). and the recognition
25 of oxidized lipids by alveolar macrophages (Dahl et al.. 2007).
6.2.4 Respiratory Symptoms and Medication Use
26 Controlled human exposure and toxicological studies have described the modes of action
27 through which short-term O3 exposure may lead to increases in respiratory symptoms by
28 demonstrating O3-induced increases in airway hyperresponsiveness, bronchoconstriction
29 (Section 6.2.2), and pulmonary inflammation (Sections 6.2.3.1 and 6.2.3.3). While
30 epidemiologic studies have not widely examined associations between ambient O3
31 exposure and airway hyperresponsiveness, they have found O3-associated increases in
32 pulmonary inflammation and oxidative stress (Section 6.3.2.2). In addition to decreases
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1 in lung function, controlled human exposure studies clearly demonstrate increases in
2 subjective respiratory symptoms including cough, pain on deep inspiration, and shortness
3 of breath (described in detail in Section 6.2.1.1). Similar to lung function responses, these
4 respiratory symptoms increase with exposure concentration, activity level of the exposed
5 individual, and duration of exposure (McDonnell et al., 1999). Increases in subjective
6 respiratory symptoms have been reported following 5.6 and 6.6 h of exposure to 60 ppb
7 O3. However, the severity of respiratory symptoms following 6.6 h of exposure to 80 ppb
8 O3 during moderate exercise is roughly 2-3 times greater than that at 60 ppb O3 (Adams.
9 2006a). These findings integrated across disciplines provide biological plausibility for
10 epidemiologic associations between increases in short-term ambient O3 exposure and
11 increases in respiratory symptoms.
12 In epidemiologic studies, respiratory symptom data typically are collected by having
13 subjects or their parents record symptoms such as wheeze, cough, and shortness of breath
14 and medication use in a diary without direct supervision by study staff. Several
15 limitations of symptom reports are we 11-recognized: recall error if not recorded daily,
16 differences among subjects in the interpretation of symptoms, biased reporting between
17 participants with and without asthma, and occurrence in a smaller percentage of the
18 population compared with changes in lung function and biological markers of pulmonary
19 inflammation. Nonetheless, symptom diaries remain a convenient and useful tool to
20 collect individual-level data from a large number of subjects and allow the modeling of
21 associations between daily changes in O3 exposure and daily changes in respiratory
22 morbidity. Importantly, most of the limitations described above are sources of random
23 measurement error that can bias effect estimates to the null or increase the uncertainty
24 around effect estimates. Furthermore, because respiratory symptoms are associated with
25 limitations in activity and function and are the primary reason for using medication and
26 seeking medical care, they provide an assessment of the clinical and public health
27 significance of ambient O3 exposure.
28 Most studies have been conducted in individuals with asthma, and as was concluded in
29 previous O3 AQCD, the collective body of epidemiologic evidence strongly supports
30 associations between increases in short-term ambient O3 exposure and increases in
31 respiratory symptoms in children with asthma (U.S. EPA. 2006b. 1996a) (Figure 6-11
32 and Table 6-19). Evidence also indicates that O3 exposure likely is associated with
33 increased use of asthma medication (Figure 6-12 and Table 6-20). Studies also find O3
34 exposure to be associated with respiratory symptoms in adults with asthma. The effects of
35 O3 exposure on respiratory symptoms in healthy populations are not as clearly indicated
36 (Figure 6-13 and Table 6-23)
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6.2.4.1 Children with Asthma
Respiratory Symptoms
1 Table 6-18 presents the characteristics and ambient O3 concentration data from studies
2 assessing associations of short-term O3 exposure with respiratory symptoms and
3 medication use in children with asthma. The strong evidence for associations between
4 ambient O3 exposure and respiratory symptoms among children with asthma is derived
5 mostly from several single-region or single-city studies (Figure 6-11 and Table 6-19).
6 Most studies of children with asthma examined 1-h max, 8-h max, or 8-h average O3
7 exposures. In U.S. multicity studies, O3 was associated with both increases and decreases
8 in respiratory symptoms among children with asthma (O'Connor et al.. 2008; Schildcrout
9 et al.. 2006; Mortimer etal.. 2002). In the NCICAS cohort (described in Section 6.2.1.2),
10 a 30-ppb increase in lag 1-4 avg of 8-h avg (10:00 a.m.-6:00 p.m.) O3 was associated with
11 an increase in morning asthma symptoms with an OR (95% CI) of 1.35 (95% CI: 1.04,
12 1.69) (Mortimer etal.. 2002). This association did not change (OR: 1.37 [95% CI: 1.02,
13 1-84]) in an analysis restricted to O3 concentrations below 80 ppb. Odds ratios for lags 2
14 and 4 of O3 exposure were similar in magntiude. In the ICAS cohort (described in Section
15 6.2.1.2), associations of 19-day avg of 24-h avg O3 with wheeze and nighttime asthma
16 were positive and negative, respectively (O'Connor et al.. 2008). NCICAS was conducted
17 during the warm season, and symptom data were collected daily (Mortimer et al.. 2002:
18 Mortimer et al.. 2000). whereas in ICAS, every 2 months, parents reported the number of
19 days with respiratory symptoms over the previous 2 weeks (O'Connor et al.. 2008).
20 Because of the two-week symptom reporting period, ICAS investigators were precluded
21 from examining associations with single-day and shorter-duration O3 exposure periods.
22 Evidence of O3-associated respiratory symptoms also was weak in another recent U.S.
23 multicity study (with cities in common with NCICAS and ICAS, Table 6-18) of 990
24 children with asthma (Schildcrout et al.. 2006). As part of the Childhood Asthma
25 Management Program, symptom data were collected daily, and analyses were restricted
26 to peak O3 periods between May and September. In meta-analyses that combined city-
27 specific estimates, a 40-ppb increase in lag 0 of 1-h max O3 was associated with any
28 asthma symptom with an OR (95% CI) of 1.08 (0.89, 1.31). Odds ratios for lags 1 and 2
29 and the 3-day sum of O3 were near 1.0. In this study, data were available from an average
30 of 12 subjects per day per city, and fewer data were collected in summer months.
31 Because O3 analyses were restricted to summer months, the fewer number of
32 observations reduced the power to detect associations for O3 relative to other pollutants,
33 which were analyzed using year-round data.
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Table 6-18 Mean and upper percentile ozone concentrations in epidemiologic
studies examining respiratory symptoms, medication use, and
activity levels in children with asthma
Study
Mortimer etal.
(2000)
Mortimer etal.
(2002)
O'Connor etal.
(2008)
Schildcroutetal.
(2006)
Gent etal.
(2003)
Thurston et al.
(1997)
Rabinovitch et
al. (2004)
Mann et al.
(2010)
Ostro et al.
(2001)
Delfino et al.
(2003)
Romieu et al.
(1996)
Romieu et al.
(1997)
Romieu et al.
(2006)
Escamilla-Nunez
etal. (2008)
Gielen etal.
(1997)
Just etal. (2002)
Jalaludin et al.
(2004)
Location
Bronx, East Harlem, NY;
Baltimore, MD; Washington, DC;
Detroit, Ml, Cleveland, OH;
Chicago, IL; St. Louis, MO
(NCI CAS)
Boston, MA; Bronx, Manhattan NY;
Chicago, IL; Dallas, TX, Seattle,
WA; Tucson, AZ
(ICAS)
Albuquerque, NM; Baltimore, MD;
Boston, MA; Denver, CO; San
Diego, CA; Seattle, WA; St. Louis,
MO; Toronto, ON, Canada (CAMP)
CT, southern MA
Connecticut River Valley, CT
Denver, CO
Fresno/Clovia, California
Los Angeles, CA
Los Angeles, CA
northern Mexico City, Mexico
southern Mexico City, Mexico
Mexico City, Mexico
Mexico City, Mexico
Amsterdam, Netherlands
Paris, France
Sydney, Australia
Years/Season
1993
Warm season
1998-2001
All-year
1994-1995
Warm season
2001
April-September
1991-1993
Warm season
1999-2002
Cold season
2000-2005
All-year
1993
August-October
1999-2000
Cold season
April-July 1991
November 1991 -
February 1992
April-July 1991
November 1991 -
February 1992
1998-2000
All-year
2003-2005
All-year
1995
Warm season
1996
April-June
1994
All-year
03
Averaging
Time
8-h avg
(10:00 a.m.-
6:00 p.m.)
24-h avg
1-h max
1-hmax
8-h rolling avg
1-hmax
1-hmax
8-h max
1-h max
1-h max
8-h max
1-h max
1-hmax
1-hmax
8-h max
1-hmax
8-h max
8-h max
24-h avg
15-h avg (6:00
a.m.-9:00
p.m.)
Mean/Median
Concentration (ppb)
48
NR
Range in medians
across cities: 43.0-65.8
58.6
51.3
83.6
28.2
49.4 (median)
Los Angeles: 59.5
Pasadena: 95.8
25.4
17.1
190
196
102
69
86.5
31.6
34.2
30.0
12
Upper Percentile
Concentrations (ppb)
NR
NR
Range in 90th across
cities: 61 .5-94.7
Max: 125.5
Max: 99.6
Max: 160
Max: 70.0
75th: 69.5, Max: 120.0
Max: 130
Max: 220
90th: 38.0, Max: 52
90th: 26.1, Max: 37
Max: 370
Max: 390
Max: 309
Max: 184
Max: 86.3 (8-h max)
Max: 56.5
Max: 61 .7
Max: 43
NCICAS = National Cooperative Inner-City Asthma Study, NR = Not Reported, ICAS = Inner City Asthma Study, NR = Not Reported, CAMP =
Childhood Asthma Management Program, Max = Maximum.
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Study Symptom
Aggregate of symptoms
Rabinovitchetal. (2004) Daytime symptoms
Delfinoetal. (2003) Bothersome symptoms
Schildcroutetal. (2006) Asthma symptoms
Gielenetal. (1997)
LRS
URS
Mortimeret al. (2002) Morning symptoms
Mortimeret al. (2000)
Romieuetal. (1996)
Romieuetal. (1997)
Individual symptoms
Jalaludinetal. (2004)
O'Connoretal. (2008)
Ostroetal.(2001)
Escamilla-Nunez et al.
(2008)
Mann etal. (2010)
Thurstonetal.(1997)
Romieuetal. (2006)
LRS
LRS
Wheeze
Wheeze/cough
Wheeze
Cough
Wheeze
Chest symptoms
Difficulty breathing
O3 Lag Subgroup
0-2 avg
0
0
0
1-4 avg All subjects
No medication
Cromolyn use
Beta-agonist/xanthine use
Steroid use —
Without allergy
With allergy
0
0
2
1-19 avg
3
0
0
All
Fungi allergic
0-5 avg GSTM1 positive
GSTM1 null
GSTP1 lie/lie Ile/Val
GSTP1 Val/Val
0.5 1 1.5
Odds ratio (95% Cl)
2.5
Figure 6-11 Associations of ambient ozone exposure with respiratory
symptoms in children with asthma. Results are presented first for
aggregate indices of symptoms then for individual symptoms.
Within each category, results generally are organized in order of
increasing mean ambient O3 concentration. LRS = lower respiratory
symptoms, URS = upper respiratory symptoms. Effect estimates
are from single-pollutant models and are standardized to a 40-, 30-,
and 20-ppb increase for 1-h max, 8-h max or 8-h avg, and 15-h avg
or 24-h avg ozone exposures, respectively.
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Table 6-19 Additional characteristics and quantitative data for studies
presented in Figure 6-11
Study
Studies examining
Location/ Population
aggregates of symptoms
Rabinovitch et al. (2004) Denver, CO
Children with asthma
Delfino et al. (2003)
Schildcroutetal. (2006
Gielenetal. (1997)
Mortimer etal. (2002)
Mortimer etal. (2QQO)
Romieu et al. (1996)
Romieu et al. (1997)
Studies examining
Jalaludin et al. (2004)
O'Connor etal. (2008)
Ostro et al. (2001)
Escamilla-Nunez et al.
(2008)
Los Angeles, CA
Children with asthma
I) 8 U.S. communities
Children with asthma
Amsterdam, Netherlands
Children with asthma
8 U.S. communities
Children with asthma
northern Mexico City,
Mexico
Children with asthma
southern Mexico City,
Mexico
Children with asthma
individual symptoms
Sydney, Australia
Children with asthma
7 U.S. communities
Children with asthma
Los Angeles, CA
Children with asthma
Mexico City, Mexico
Children with asthma
03
Lag
0-2
avg
0
0
0
1-4
avg
0
0
2
1-19
avg
3
0
03
Averagin
gTime
1 -h max
1 -h max
1 -h max
8-h max
8-h avg
(10:00 a.m.-
6:00 p.m.)
1 -h max
1 -h max
1 5-h avg
(6:00 a.m.-
9:00 p.m.)
24-h avg
1 -h max
1 -h max
Symptom Subgroup
Daytime
symptoms
Bothersome
symptoms
Asthma symptoms
LRS
URS
Morning All subjects
symptoms No medication use
Cromolyn use
p-agonist/xanthine use
Steroid use
Without allergy
With allergy
LRS
LRS
Wheeze
Wheeze/cough
Wheeze
Wheeze
Odds Ratio
(95% Cl)a
1.34(1.01,1.77)
1 .09 (0.39, 3.03)
1.08(0.89,1.31)
1.04(0.75,1.45)
1.16(1.02,1.32)
1.35(1.04,1.74)
1.08(0.62,1.87
2.13(1.12,4.04)
1.39(0.98,1.98)
1.17(0.79,1.72)
1.59(1.00,2.52)
1.35(0.92,1.96)
1.07(1.02,1.12)
1.09(1.04,1.14)
1.21 (0.92,1.59)
1.02(0.86,1.21)
0.94(0.88,1.00)
1.08(1.03,1.14)
Mann et al. (2010)
Thurston et al. (1997)
Romieu et al. (2006)
Fresno/Clovia, California
Children with asthma
CT River Valley, CT
Children with asthma
Mexico City, Mexico
Children with asthma
0
0
0-5
avg
8-h max
1 -h max
1-h max
Wheeze
Chest symptoms
Difficulty breathing
All
Fungi allergic
GSTM1 sufficient
GSTM1 null
GSTP1 lie/lie NeA/al
GSTP1 ValA/al
1 .00 (0.84,
1 .06 (0.84,
1.28(1.10,
1.10(0.98,
1.17(1.02,
1 .06 (0.94,
1.30(1.10,
1.19)
1.34)
1.50)
1.24)
1.33)
1.20)
1.53)
LRS = Lower respiratory symptoms, URS = Upper respiratory symptoms.
"Effect estimates are standardized to a 40, 30, and 20 ppb increase for 1-h max, 8-h max or 8-h avg, and 15-h avg or 24-h avg 03, respectively.
4
5
Several longitudinal studies conducted in multiple cohorts of children with asthma in
Mexico City, Mexico examined 1-h max O3 exposures and found associations with
increases in respiratory symptoms (Escamilla-Nunez et al.. 2008; Romieu et al.. 2006;
Romieu etal.. 1997; Romieu etal.. 1996). Recent studies expanded on earlier evidence
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1 by indicating associations with multiday averages of O3 exposure. Both Romieu et al.
2 (1996) and Romieu et al. (1997) found that among single-day 1-h max O3 exposures
3 lagged 0 to 2 days, lag 0 had the greatest estimated effect on respiratory symptoms.
4 Romieu et al. (2006) and Escamilla-Nunez et al. (2008) found that the magnitudes of
5 association of ambient 1-h max O3 exposure with respiratory symptoms and medication
6 use increased with increasing number of days over which O3 exposure was averaged.
7 Studies of children with asthma also identified factors that may contribute to
8 heterogeneity in symptom responses to ambient O3 exposure. Multiple studies, all of
9 which examined 8-h avg (10:00 a.m.-6:00 p.m.) or 8-h max O3 exposures, found larger
10 associations among subjects taking asthma medication; however, the medications varied
11 among studies. Consistent with findings for lung function, in the NCICAS multicity
12 cohort, larger associations for morning symptoms were observed in children taking
13 cromolyn (used to treat asthma with allergy) or beta-agonists/xanthines than in children
14 taking no medication. Odds ratios did not differ as much between children taking steroids
15 and children taking no medication (Figure 6-11 and Table 6-19) (Mortimer et al.. 2000).
16 In a cohort of children with asthma in Southern New England, O3 exposures were
17 associated with larger increases in chest tightness among children taking maintenance
18 medication (i.e., steroids, cromolyn, or leukotriene inhibitors).
19 Most studies of children with asthma reported that a majority of subjects (52 to 100%)
20 were atopic as determined by a positive skin prick test to any examined allergen;
21 however, results did not conclusively indicate that children with asthma and atopy were
22 more susceptible to the effects of O3 exposure. In the multicity NCICAS cohort,
23 Mortimer et al. (2000) found that O3 was associated with a similar incidence of asthma
24 symptoms among the 79% of subj ects with atopy and the 21 % of subj ects without atopy
25 (Figure 6-11 and Table 6-19). Odds ratios did not differ by residential levels of allergens.
26 In a recent study of children with asthma in Fresno, CA, most associations of single- and
27 multiday lags of 8-h max O3 exposure (0-14 days) with wheeze were near or below 1.0
28 (Mann et al.. 2010). The estimated effects did not differ in fungi allergic subjects, A
29 larger association was found for cat allergic subjects; however, this finding was limited to
30 O3 exposure lagged 14 days. In this study, many subjects were allergic to multiple
31 allergens; however, associations were not compared between subjects with any versus no
32 allergic sensitization.
33 Although Romieu et al. (2006) did not observe differences in associations between O3
34 and lung function by GST genetic polymorphisms (Section 6.2.1.2), they did observe
35 effect modification for respiratory symptoms. Compared with GSTM1 positive subjects
36 and GSTP1 lie/lie or Ile/Val subjects, larger effects were estimated for GSTM1 null
37 subjects and for GSTP1 Val/Val subjects, respectively (Figure 6-11 and Table 6-19).
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1 Ozone had the greatest estimated effect on difficulty breathing in children with asthma
2 who were both GSTM1 null and GSTP1 Val/Val (OR: 1.49 [95% CI: 1.14, 1.93] per 30-
3 ppb increase in lag 0-5 avg of 8-h max O3). In the same cohort of children, antioxidant
4 supplementation reduced O3-associated increases in airway inflammation (Sienra-Monge
5 et al.. 2004). These results add to the body of epidemiologic evidence that antioxidant
6 capacity influences risk of O3-related respiratory morbidity. As was discussed in Section
7 6.2.1.2, compared with the GSTM1 genotype, evidence for effect modification by GSTP1
8 genetic polymorphisms is less certain. Romieu et al. (2006) found that the GSTP1
9 Val/Val variant was associated with a lesser O3-associated decrement in lung function but
10 greater risk of respiratory symptoms. Whereas some studies have reported greater risk of
11 asthma among GSTP1 lie/lie or Ile/Val subjects (Mapp et al., 2002; Hemmingsen et al..
12 2001). others have reported greater risk among GSTP1 Val/Val subjects (Tamer et al..
13 2004). In Romieu et al. (2006). GSTP1 lie/lie was associated with greater severity of
14 asthma, and Lee et al. (2004b) also reported greater risk of air pollution-associated
15 asthma among GSTP1 lie/lie children in the Southern California Children's Health
16 Study.
Asthma Medication Use
17 The 2006 O3 AQCD concluded that ambient O3 likely was associated with increased
18 asthma medication use based on the positive associations found in several studies of
19 children with asthma (Figure 6-12 and Table 6-20). Among the few newly available
20 studies on asthma medication use, evidence generally supported the previous conclusion
21 (Escamilla-Nunez et al.. 2008; Romieu et al.. 2006). Most of these studies examined lags
22 0 or 1 of 1-h max O3 exposures; however, Romieu et al. (2006) found that lag 0-5 avg of
23 1-h max O3 was associated with a larger increase in bronchodilator use than were lags 1
24 or 0-1 avg. As compared with respiratory symptoms, effects on medication use were
25 estimated with greater uncertainty as indicated by the wide 95% CIs. The wide 95% CIs
26 have been attributed to a smaller number of study subjects reporting medication use and
27 the low frequency of use over the study period. However, within most studies, findings
28 were similar for respiratory symptoms and asthma medication use. Among recent studies,
29 Romieu et al. (2006) and Escamilla-Nunez et al. (2008) observed O3-associated increases
30 in both respiratory symptoms and bronchodilator use. Schildcrout et al. (2006) did not
31 observe O3-associated increases in either respiratory symptoms or rescue inhaler use. In
32 contrast, Romieu et al. (1996) and Rabinovitch et al. (2004) observed that O3 was
33 positively associated with daytime respiratory symptoms but not with bronchodilator use.
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Study Outcome
Schildcroutetal. (2006) Rescueinhaleruse
Thurstonetal. (1997) Beta-agonist use
Ostro et al. (2001) Extra medication use
Romieuetal. (1996)
Romieuetal. (1997)
Romieuetal. (2006)
Gielenetal. (1997)
Bronchodilatoruse
Bronchodilatoruse
Bronchodilatoruse
Lag Subgroup
0
0
1
0
0
0-5 avg GSTP1 lie/lie Ile/Val
GSTP1 Val/Val
Bronchodilatoruse
Jalaludinetal. (2004) Beta-agonist use/no steroid 1
CSuse
0 0.5 1 1.5
Odds ratio (95% Cl)
CS = corticosteroid. Results are presented in increasing order of ambient ozone concentration. Effect estimates are from single-
pollutant models and are standardized to a 40- and 30-ppb increase for 1-h max and 8-h max ozone, respectively
Figure 6-12 Associations of ambient ozone exposure with asthma medication
use.
Table 6-20
Study
Schildcroutetal.
(2006)
Thurston et al.
(1997)
Ostro et al. (2001)
Romieu et al. (1996)
Romieu et al. (1997)
Romieu et al. (2006)
Gielenetal. (1997)
Jalaludin et al.
(2004)
Additional characteristics and quantitative data for studies
presented in Figure 6-12
Location/
Population
8 U.S. communities
Children with asthma
CT River Valley, CT
Asthmatic campers
Los Angeles, CA
Children with asthma
northern Mexico City, Mexico
Children with asthma
southern Mexico City, Mexico
Children with asthma
Mexico City, Mexico
Children with asthma
Amsterdam, Netherlands
Children with asthma
Sydney, Australia
Children with asthma
O3Lag
0
0
1
0
0
0-5 avg
0
1
03
Averaging
Time
1-h max
1-h max
1-h max
1-h max
1-h max
1-h max
8-h max
1-h max
Medication Subgroup
Rescue inhaler use
Beta-agonist use
Extra medication use
Bronchodilator use
Bronchodilator use
Bronchodilator use GSTP1 lie/lie
Ile/Val
GSTP1 ValA/al
Bronchodilator use
Beta-agonist use/no
steroid
ICS use
Odds Ratio
(95% Cl)a
1.01 (0.89,1.15)
1.17(0.96,1.44)
1.10(1.03,1.19)
0.97(0.93,1.01)
1.02(1.00,1.05)
0.96 (0.90, 1 .02)
1.10(1.02,1.19)
1.10(0.78,1.55)
1 .08 (0.89, 1 .32)
1.08(0.96,1.21)
CS= Corticosteroid.
"Effect estimates are standardized to a 40- and 30-ppb increase for 1-h max and 8-h max 03, respectively.
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Changes in Activity
1 While investigation has been limited, evidence has not consistently indicated associations
2 between O3 exposure and diminished activity level in children with asthma (O'Connor et
3 al.. 2008; Delfino et al.. 2003). These studies have examined a range of O3 averaging
4 times and lags of exposure. In the large ICAS cohort, O'Connor et al. (O'Connor et al..
5 2008) found that a 20-ppb increase in lag 1-19 avg of 24-h O3 ambeint was associated
6 with a 10% lower odds (95% CI: -26, 10) of slow play. In a small (n = 22) panel study
7 conducted in children with asthma in Los Angeles CA, Delfino et al. (2003) found that a
8 40-ppb increase in lag 0 of 1-h max O3 was associated with an increase in symptoms that
9 interfered with daily activity with an OR (95% CI) of 7.41 (1.18, 43.2). Several studies
10 reported increases in school absenteeism in children with asthma in association with long
11 lags of O3 exposure (14-day and 30-day distributed lags or 19-day avg) (O'Connor et al..
12 2008; Gilliland et al.. 2001; Chen etal.. 2000). Whereas Chen et al. (2000) and O'Connor
13 et al. (2008) examined absences for any reason, Gilliland et al. (2001) found associations
14 with absences for respiratory illnesses. Despite this evidence, several limitations have
15 been noted, including the uncertain biological relevance of long lag periods of O3
16 exposure and the potential for residual seasonal confounding when examining long lag
17 periods of exposure. In analyses of single-day lags, Gilliland et al. (2001) found that 8-h
18 avg (10:00 a.m.-6:00 p.m.) O3 exposure was associated with increases in respiratory-
19 related absences from lag day 1 to lag day 5, indicating an effect of exposures with
20 shorter lag periods.
6.2.4.2 Adults with Respiratory Disease
21 Characteristics and ambient O3 concentration data from studies of adults with respiratory
22 disease are presented in Table 6-21. In this relatively small body of literature, several
23 studies found ambient O3 exposure (1-h max or 8-h max) to be associated with increases
24 respiratory symptoms and decreases in activity levels in adults with asthma (Khatri et al..
25 2009: Feo Brito et al.. 2007: Eiswerth et al.. 2005: Ross et al.. 2002). In a recent panel
26 study of adults with COPD, investigators found lag 1 of 8-h max O3 to be associated with
27 increased odds of dyspnea and sputum changes but decreased odds of nasal discharge,
28 wheeze, or upper respiratory symptoms (Peacock et al.. 2011).
29 In a panel study of children and adults with asthma, lag 1-3 avg of 8-h max O3 exposure
30 was associated with increases in morning and evening symptom scores and frequency of
31 asthma medication use (Ross et al.. 2002). During one pollen season (May-June 2000 or
32 2001), Feo Brito et al. (2007) specifically followed a group of 137 adults who had asthma
33 and pollen allergy in central Spain. In the industrial Puertollano, a 40-ppb increase in lag
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1
2
3
4
5
6
7
3 of 1-h max O3 was associated with a 14.3% increase (95% CI: 3.6, 26.0) in the number
of subjects reporting respiratory symptoms, adjusting only for time trend. There was a
much weaker association in the less industrialized Ciudad Real with lower ambient air
pollution concentrations and a narrower range of ambient O3 concentrations (2.3% [95%
CI: -14, 21%] per 40-ppb increase in lag 4 of 1-h max O3). Park et al. (2005a) followed
adults with asthma in Korea during a period that included dust storms and found that a
20-ppb increase in lag 0 of 24-h avg O3 was associated with an increased odds of night
symptoms (OR: 1.11 [95% CI: 0.96, 1.29]) but not cough (OR: 1.00 [95% CI: 0.94,
1.06]) or rescue inhaler use (OR: 0.99 [95% CI: 0.94, 1.05]).
Table 6-21
Study
Khatrietal.
(2009)
Ross et al.
(2002)
Eiswerth et al.
(2005)
Peacock etal.
(2Q11)
Feo Brito etal.
(2007)
Wiwatanadate et
al. (2011)
Park etal.
(2005a)
Mean and upper percentile ozone concentrations in epidemiologic
studies examining respiratory symptoms and medication use in
adults with respiratory disease
Location
Atlanta, GA
East Moline, IL
Glendora, CA
London, England
Ciudad Real and
Puertollano, Spain
Chiang Mai, Thailand
Incheon, Korea
Years/Season
2003, 2005, 2006
Warm season
April-October 1994
1983
Cold season
1995-1997
All-year
2000-2001
Warm season
August 2005-June
2006
March-June 2002
« °'-
Averaging
Time
8-h max
8-h avg
1-h max
8-h max
1-h max
24-h avg
24-h avg
Mean/Median
Concentration (ppb)
59a
41.5
NR
15.5
65.9 (Ciudad Real)"
56.8 (Puertollano)"
17.5
Dust event days: 23.6
Control days: 25.1
Upper Percentile
Concentrations (ppb)
Max: 73a
Max: 78.3
NR
Autumn/Winter Max: 32
Spring/Summer Max: 74
Max: 101.5" (Ciudad Real);
70.5b (Puertollano)
90th: 26.82
Max: 34.65
NR
NR = Not Reported, Max = Maximum.
'Individual-level exposure estimates were derived based on time spent in the vicinity of various 03 monitors.
"Concentrations converted from ug/m3 to ppb using the conversion factor of 0.51 assuming standard temperature (25°C) and pressure (1 atm).
10 Studies also indicated that ambient O3 exposure may result in decreases in activity levels
11 in adults with asthma. Notably, although conducted over single seasons, these studies did
12 not consider confounding by meteorological factors. In a cross-sectional summer study in
13 Atlanta, GA (described in Section 6.2.1.2), Khatri et al. (2009) observed that a 30-ppb
14 increase in lag 2 of 8-h max O3 was associated with a 0.69-point decrease (95% CI: -1.28,
15 -0.11) in the Juniper quality of life score, which incorporates indices for symptoms,
16 mood, and activity limitations (7-point scale). In a fall study conducted in the Los
17 Angeles, CA area, Eiswerth et al. (2005) examined the activities of 64 individuals with
18 asthma (age 16 years and older). A 40-ppb increase in 1-h max O3 was associated with a
19 0.24% (95% CI: 0.08, 0.40%) lower probability of participation in indoor activities. The
20 association with outdoor activities was positive but not statistically significant. Although
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1
2
3
4
5
6
the authors acknowledged that their findings were unexpected and may have been
influenced by lack of control for potential confounders, they interpreted the decrease in
indoor activities as rest replacing chores. In contrast, in a panel study of individuals with
asthma (ages 13-78 years) in Thailand, O3 exposure was associated with a lower odds of
symptoms that interfered with activities (OR: 0.74 [95% CI: 0.57, 0.96] per 20-ppb
increase in lag 4 of 24-h avg O3) (Wiwatanadate and Liwsrisakun. 2011).
9
10
11
12
6.2.4.3 Populations not Restricted to Individuals with Asthma
Characteristics and ambient O3 concentration data from studies of populations not
restricted to individuals with asthma are presented in Table 6-22. In contrast with
findings for lung function (Section 6.2.1.2), epidemiologic studies do not provide
consistent evidence of associations between short-term ambient O3 exposure and
increases in respiratory symptoms in children without asthma (Figure 6-13 and
Table 6-23).
Table 6-22 Mean and upper percentile ozone concentrations in epidemiologic
studies examining respiratory symptoms in populations not
restricted to individuals with asthma
Study
Apteetal.
(2008)
Neas et al.
(1995)
Triche et al.
(2006)
Linn et al. (1996)
Gold et al.
(1999)
Wardetal.
(2002)
Hoekand
Brunekreef
(1995)
Moon et al.
(2009)
Rodriguez etal.
(2007)
Location
Multiple U.S. cities (NR)
Uniontown, PA
Southwestern VA
Rubidoux, Upland,
Torrence, CA
Mexico City, Mexico
Birmingham and Sandwell,
England
Deurne and Enkhuizen,
Netherlands
4 cities, South Korea
Perth, Australia
Years/Season
1994-1998
Winter or summer
June-August 1990
1995-1996
Warm season
1992-1993,1993-1994
Fall and spring
1991
Winter, spring, fall
1997
Winter and summer
1989
March-July
April-May 2003
1996-2003
All-year
Os Averaging
Time
Workday avg
(8:00 a.m. -
5:00 p.m.)
24-h avg
12-h avg (8:00
a.m.-8:00p.m.)
1-h max
8-h max
24-h avg
24-h avg
24-h avg
24-h avg
1-h max
8-h avg (10:00
a.m.-6:00p.m.)
1-h max
24-h avg
Mean/Median
Concentration (ppb)
34.2a
25.5a
37.2
60.8
54.5
35.2
23
52.0
Winter median: 13.0
Summer median: 22.0
Deurne: 57
Enkhuizen: 59
NR
33
28
Upper Percentile
Concentrations (ppb)
Max: 86.2a
Max: 67.3a
Max: 44.9
75th: 70.0, Max: 95.0
75th: 64.1, Max: 87.6
75th: 40.6, Max: 56.6
Max: 53
Max: 103
Winter Max: 33
Summer Max: 41
Max: 107
Max: 114
NR
Max: 95
Max: 74
NR = Not Reported, Max = Maximum.
'Concentrations converted from ug/m3 to ppb using the conversion factor of 0.51 assuming standard temperature (25°C) and pressure (1 atm).
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1 Among healthy children in Uniontown, PA, Neas et al. (1995) found a stronger
2 association between O3 exposure and evening cough using ambient concentrations
3 weighted by time spent outdoors (OR: 2.20 [95% CI: 1.02, 4.75] per 30-ppb increase in
4 lag 0 of 12-h avg [8:00 a.m.-8:00 p.m.]) than using unweighted concentrations (OR: 1.36
5 [95% CI: 0.86, 2.13] per 30-ppb increase in lag 0 of 12-h avg [8:00 a.m.-8:00 p.m.]).
6 Several other panel studies of school-aged children, in which asthma prevalence ranged
7 between 0 to 50%, reported null or negative associations between various averaging
8 times and lags of ambient O3 exposure and respiratory symptoms (Moon et al.. 2009;
9 Rodriguez et al.. 2007; Ward et al.. 2002; Linn et al.. 1996; Hoek and Brunekreef. 1995).
10 For example, a large study of 696 children in four regions in South Korea, Moon et al.
11 (2009) observed that among all subjects, ORs of lag 0 8-h avg O3 with most respiratory
12 symptoms were close to 1.0. In city-specific analyses, O3 exposure was only consistently
13 associated with increases in URS (runny nose or sneezing), with the largest magnitude of
14 association observed in Jeju island (OR: 1.08 [95% CI: 0.96, 1.21] per a 30-ppb increase
15 in lag 0 8-h avg O3). Consistent with other studies conducted in Mexico City, Gold et al.
16 (1999) reported a positive association between lag 1 of 24-h avg O3 exposure and phlegm
17 in children; however, investigators acknowledged being unable to distinguish between
18 the effects of the highly-correlated O3 and PM10 (r = 0.75).
19 In a recent study, O3 exposure was associated with increased odds of respiratory
20 symptoms in a group of infants who have mothers with asthma (Triche et al.. 2006).
21 Triche et al. (2006) followed 691 infants in southwestern VA Yfor 83 days between June
22 and August of 1995 and/or 1996 and found that a 20-ppb increase in lag 0 of 24-h avg O3
23 was associated with odds ratios (95% CI) of 2.34 (1.02, 5.37) for wheeze and of 3.63
24 (1-81, 7.28) for difficulty breathing among the 61 infants who had mothers with asthma.
25 Investigators estimated smaller magnitudes of association for 1-h and 8-h max O3
26 exposures. Smaller, statistically nonsignificant associations also were found in analyses
27 that included all subjects (Figure 6-13 and Table 6-23). While these results suggested that
28 children with mothers with asthma may be at greater risk of O3-related respiratory
29 morbidity, the authors acknowledged that mothers with asthma may be more likely to
30 report symptoms in their children and that transient wheeze, which is common in infants,
31 may not predict respiratory morbidity later in life. In a study of children with parental
32 history of asthma with follow-up to an older age (5 years), ambient O3 exposure was not
33 associated with increases in respiratory symptoms (Rodriguez et al.. 2007).
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Study
O3 Lag Symptom
Subgroup
Ward et al. (2002) 0-6 avg Wheeze
Shortness of breath
Hoekand Brunekreef
0
(1995)
Moon etal. (2009) 0
Neasetal. (1995) 0
Tricheetal. (2006) 0
Gold etal. (1999)
0
Cough
URS
URS
Evening cough
Wheeze
Phlegm
All subjects
Jeju Island
All
With asthmatic mothers
0123
Odds ratio (95% Cl)
LRS = lower respiratory symptoms, URS = Upper respiratory symptoms. Effect estimates are from single-pollutant models and are
standardized to a 40-, 30-, and 20-ppb increase for 1-h max, 8-h max or 12-h avg, and 24-h avg ozone exposures, respectively.
Figure 6-13 Associations of ambient ozone exposure with respiratory
symptoms in studies not restricted to children with asthma.
Table 6-23
Study
Ward etal. (2002)
Hoekand
Brunekreef (1995)
Moon et al. (2009)
Neas et al. (1995)
Triche et al. (2006)
Gold et al. (1999)
Additional characteristics and
presented in Figure 6-13
Location/
Population
Birmingham and
Sandwell, England
Children
Enkhuizen,
Netherlands
Children
4 cities, South Korea
Children
Uniontown, PA
Healthy children
southwestern VA
Infants
Mexico City, Mexico
Children
0 Laa °3 Averaging
°3 Lag Time
0-6 avg 24-h avg
0 1-h max
0 24-h avg
0 12-h avg (8:00
a.m.-8:00 p.m.)
0 8-h max
1 24-h avg
quantitative data for studies
Symptom
Wheeze
Shortness of breath
Cough
URS
URS
Evening cough
Wheeze
Phlegm
Subgroup «*«•
0.78 (0.22, 2.79)
1 .80 (0.64, 5.06)
0.95(0.71,1.25)
1.15(1.00,1.33)
All subjects 0.96(0.90,1.03)
Jeju Island 1.08(0.96,1.21)
2.20(1.02,4.75)"
All subjects 1 .60 (0.85, 3.00)
Maternal asthma 2.34 (1 .02, 5.37)
1.04(1.00,1.07)
LRS = Lower respiratory symptoms, URS = Upper respiratory symptoms
'Effect estimates are standardized to a 40-, 30-, and 20-ppb increase for 1 -h max, 8-h max or 12-h avg, and 24-h avg 03, respectively.
b03 exposures were weighted by the proportion of time spent outdoors.
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1 A recent cross-sectional study examined 4,200 adult workers from 100 office buildings
2 across the U.S. and found that a range of ambient O3 exposure metrics, including the 24-
3 h, workday (8:00 a.m.-5:00 p.m.), and late workday (3:00 p.m.-6:00 p.m.) averages, were
4 associated with increases in building-related URS (nasal congestion or sore throat) and
5 LRS (wheeze, shortness of breath, or chest tightness) (Apte et al., 2008). Investigators
6 suggested that the findings may have been attributable to formaldehyde and organic acids
7 produced from O3-initiated reactions within buildings; however, additional data on indoor
8 levels of volatile organic compounds, indoor O3, and infiltration rates is warranted to
9 characterize whether the observed associations were attributable to the formation of these
10 secondary species by ambient O3 penetrating indoors
6.2.4.4 Confounding in Epidemiologic Studies of Respiratory
Symptoms and Medication Use
11 Epidemiologic studies did not indicate that associaitons between short-term O3 exposure
12 and respiratory symptoms were confounded by meteorological factors. Except where
13 specified in the text, associations between ambient O3 exposure and respiratory
14 symptoms or medication use were found after adjusting for temperature in models. Some
15 studies additionally included humidity in models (Triche et al., 2006; Ross et al., 2002) or
16 found no independent association with respiratory symptoms (Thurston et al.. 1997).
17 Several studies that examined populations with a high prevalence of atopy found O3-
18 associated increases in respiratory symptoms and asthma medication use in copollutant
19 models that included daily pollen counts (Just et al.. 2002; Ross et al.. 2002; Gielen et al..
20 1997). Gielen et al. (1997) and Ross et al. (2002) specifically reported a high prevalence
21 of grass pollen allergy in their study populations (52% and 38%, respectively). Ross et al.
22 (2002) found similar associations of O3 with morning symptoms and asthma medication
23 use in a single-pollutant model (e.g., 0.21-point [95% CI: 0.12, 0.30] increase in
24 symptom score per 30-ppb increase in lag 1-3 avg of 8-h max O3) and in a copollutant
25 model with daily pollen counts (e.g., 0.20-point [95% CI: 0.11, 0.29] increase in
26 symptom score per 30-ppb increase in lag 1-3 avg of 8-h max O3). Feo Brito et al. (2007)
27 specifically followed a group of adults in central Spain, all of whom had both asthma and
28 pollen allergy. In one city, O3 was associated with an increase in the number of subjects
29 reporting symptoms. A smaller, statistically nonsignificant effect estimate was obtained
30 for pollen. Conversely, in another city, pollen was associated with an increased incidence
31 of respiratory symptoms, whereas O3 was not. While copollutant modeling was not
32 conducted, in both locations, O3 and pollen concentrations were weakly correlated,
33 indicating that the findings for O3 were not likely confounded by pollen. Rather, the
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1
2
results suggested that O3 and pollen may have independent effects that vary between
locations, depending on the mix of airborne pollutants.
Table 6-24
Study
Mortimer etal. (2002)
Thurston et al. (1997)
Romieu et al. (1996)
Romieu et al. (1997)
Associations between short-term ozone exposure and
symptoms in single- and copollutant models
Location/
Population
Bronx, East Harlem,
NY; Baltimore, MD;
Washington, DC;
Detroit, Ml, Cleveland,
OH; Chicago, IL; St.
Louis, MO (NCICAS)
Children with asthma
CT River Valley
Children with asthma
attending summer
camp
Mexico City, Mexico
Children with asthma
Mexico City, Mexico
Children with asthma
O3 Exposure Data Symptom
8-h avg (10:00 a.m.- Morning symptoms
6:00 p.m.)
Lag 1-4 avg
1 -h max Chest symptoms
LagO
1-hmax LRS
LagO
1-hmax LRS
LagO
Os-associated OR
in Single-Pollutant
Model (95% Cl)a
8 cities with S02 data
1.35(1.04,1.69)
7 elites with N02 data
1.25(0.94,1.67)
3 cities with PMiodata
1.21 (0.61,2.41)
1.21 (1.12,1.31)"
1.07(1.02,1.12)
1.09(1.04,1.14)
respiratory
Os-associated OR
in Copollutant
Model (95% Cl)a
with lag 1-2 avg, 3-h
avg S02
1.23(0.94,1.61)
with lag 1-6 avg, 24-h
avg N02
1.14(0.85,1.55)
with lag 1-2 avg, 24-h
avg PM10
1 .08 (0.49, 2.39)
with lag 0, 12-havg
sulfate
1.19(1.06,1.35)"
with lag 0, 24-h avg
PM2.5
1.06(1.02,1.10)
with lag 0, 24-h avg
PM10
1.09(1.01,1.19)
LRS = Lower respiratory symptoms.
"Effect estimates are standardized to a 40- and 30-ppb increase for 1-h max and 8-h avg O3, respectively.
"Temperature not included in models.
3 Robust associations between O3 exposure and respiratory symptoms also were observed
4 in copollutant models that included PM2 5, PM10, sulfate, SO2, or NO2 (Table 6-24).
5 Information on confounding in asthma medication use associations was more limited.
6 The association between O3 and bronchodilator use did not change in Gent et al. (2003)
7 after adjusting for PM2 5 but decreased in magnitude in Thurston et al. (1997) after
8 adjusting for 12-h avg sulfate. For respiratory symptoms and medication use, copollutant
9 associations remained robust after adjusting for O3. Notably, studies examined different
10 averaging times for O3 (1-h max or 8-h avg) and co-pollutants (3-h to 24-h avg) and
11 reported a range of correlations with co-pollutants. Two studies conducted concurrently
12 in two regions of Mexico City examined lag 0 exposures of 1-h max O3 and 24-h avg
13 PM10 or PM25 and found robust associations with respiratory symptoms for both O3 and
14 co-pollutants (Romieu et al.. 1997; Romieu et al.. 1996). Romieu et al. (1997) reported a
15 moderate correlation between 1-h max O3 and 24-h avg PM10 (r = 0.47). Thurston et al.
16 (1997) and Gent et al. (2003) found 1-h max O3 concentrations to be highly correlated
17 with 12-h avg sulfate (r=0.74) and 24-h avg PM25 (r=0.77), respectively, thus the
18 copollutant results should be interpreted with caution. The association between O3
19 exposure and respiratory symptoms observed in NCICAS was robust in two-pollutant
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1 models with SO2, NO2, and or PM10; however, the interpretation is complicated because
2 of the different averaging times and lags of exposure examined for O3 and co-pollutants
3 (Mortimer et al., 2002) (Table 6-24). Also difficult are interpretations of the robust
4 associations observed between ambient O3 exposure and respiratory symptoms after
5 adjusting for multiple pollutants (i.e., PM2 5 plus NO2 or PM10.2 5) (Escamilla-Nunez et al.
6 2008: Tricheetal.. 2006).
6.2.4.5 Summary of Epidemiologic Studies of Respiratory
Symptoms and Asthma Medication Use
7 With a majority of investigation focused on individuals with asthma, the collective
8 epidemiologic evidence clearly demonstrates that short-term ambient O3 exposure is
9 associated with increases in respiratory symptoms and asthma medication use in children
10 with asthma. In a smaller body of literature, several studies find associations in adults
11 with asthma. In comparison, evidence has not consistently indicated that short-term O3
12 exposure is associated with reduced activity levels in children or adults with asthma.
13 Although O3 exposure has been associated with school absenteeism among children with
14 asthma, only Gilliland et al. (2001) examined absences specifically for respiratory causes
15 and found associations with O3 exposure lag periods shorter than 14 days. Epidemiologic
16 studies do not provide consistent evidence of association between short-term ambient O3
17 exposure and respiratory symptoms in children without asthma.
18 Collectively, epidemiologic studies most frequently examined 1-h max and 8-h max or
19 avg O3 exposures, and the few studies that examined both averaging times found similar
20 magnitudes of associations with respriatory symptoms (Triche et al., 2006; Delfino et al.,
21 2003; Gent et al.. 2003). Several studies found increases in respiratory symptoms with O3
22 exposures averaged over 12 to 24 hours (Triche et al., 2006; Jalaludin et al., 2004; Gold
23 et al.. 1999; Neas et al.. 1999). Epidemiologic studies examined associations of
24 respiratory symptoms with single-day O3 concentrations lagged from 0 to 5 days as well
25 concentrations averaged over 2 to 19 days. While O3 exposures lagged 0 or 1 days were
26 consistently associated with respiratory symptoms, several studies that examined a range
27 of exposure lags found larger effect estimates for multiday averages (3- to 6-days) of O3
28 exposure (Escamilla-Nunez et al.. 2008; Romieu et al.. 2006; Rabinovitch et al.. 2004;
29 Just et al., 2002; Mortimer et al., 2002; Ross et al.. 2002). These epidemiologic findings
30 are in contrast with those from controlled human exposure studies that find attenuated
31 symptom responses with O3 exposures repeated over several days (Section 6.2.1.1). The
32 epidemiologic findings for lagged O3 exposures or those accumulated over several days
33 are well-supported by the action of O3 to sensitize bronchial smooth muscle to
34 hyperreactivity, thus acting as a primer for subsequent exposure to antigens such as
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1 allergens (Section 5.3.5). In several of the studies of individuals with asthma, the
2 prevalence of atopy was high (50-100%), and sensitization of airways provides a
3 biologically plausible mode of action by which lagged or multiday average O3 exposures
4 are associated with increases in respiratory symptoms in these studies of individuals with
5 asthma.
6 Epidemiologic evidence did not indicate that associations between short-term O3
7 exposure and respiratory symptoms were confounded by temperature or pollen. In the
8 limited analysis of confounding by co-pollutants (primarily PM), robust associations with
9 respiratory symptoms were observed for O3; however, disentangling the independent
10 effects of O3 exposure in many studies is complicated due to the high correlations
11 observed between O3 and PM, different averaging times and lags of exposure examined
12 for co-pollutants, and the multiple co-pollutants included in models. Nonetheless, the
13 consistency of association among individuals with asthma with and without adjustment
14 for copollutant exposures combined with evidence from controlled human exposure
15 studies for the direct effect of O3 exposure provide substantial evidence for the
16 independent effects of ambient O3 exposure on increases in respiratory symptoms
6.2.5 Lung Host Defenses
17 The mammalian respiratory tract has a number of closely integrated defense mechanisms
18 that, when functioning normally, provide protection from the adverse effects of a wide
19 variety of inhaled particles and microbes. For simplicity, these interrelated defenses can
20 be divided into two major parts: (1) nonspecific (transport, phagocytosis, and bactericidal
21 activity) and (2) specific (immunologic) defense mechanisms. A variety of sensitive and
22 reliable methods have been used to assess the effects of O3 on these components of the
23 lung's defense system to provide a better understanding of the health effects associated
24 with the inhalation of this pollutant. The previous O3 AQCD states that animal
25 toxicological studies provide extensive evidence that acute O3 exposures as low as 0.08 to
26 0.5 ppm can cause increases in susceptibility to infectious diseases due to modulation of
27 lung host defenses. Tables 6-6 through 6-9, beginning on p. 6-41 of the 1996 O3 AQCD
28 (U.S. EPA. 1996a). and Table AX5-7, beginning on p. AX5-8 of the 2006 O3 AQCD
29 (U.S. EPA. 2006b). present studies on the effects of O3 on host defense mechanisms. This
30 section discusses the various components of host defenses, such as the mucociliary
31 escalator, the phagocytic, bactericidal, and regulatory role of the alveolar macrophages
32 (AMs), the adaptive immune system, and integrated mechanisms that are studied by
33 investigating the host's response to experimental pulmonary infections.
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6.2.5.1 Mucociliary Clearance
1 The mucociliary system is one of the lung's primary defense mechanisms. It protects the
2 conducting airways by trapping and quickly removing material that has been deposited or
3 is being cleared from the alveolar region by migrating alveolar macrophages. Ciliary
4 movement directs particles trapped on the overlying mucous layer toward the pharynx,
5 where the mucus is swallowed or expectorated.
6 The effectiveness of mucociliary clearance can be determined by measuring such
7 biological activities as the rate of transport of deposited particles; the frequency of ciliary
8 beating; structural integrity of the ciliated cells; and the size, number, and distribution of
9 mucus-secreting cells. Once this defense mechanism has been altered, a buildup of both
10 viable and nonviable inhaled substances can occur on the epithelium and may jeopardize
11 the health of the host, depending on the nature of the uncleared substance. Impaired
12 mucociliary clearance can result in an unwanted accumulation of cellular secretions,
13 increased infections, chronic bronchitis, and complications associated with chronic
14 obstructive pulmonary disease. A number of previous studies with various animal species
15 have examined the effect of O3 exposure on mucociliary clearance and reported
16 morphological damage to the cells of the tracheobronchial tree from acute and sub-
17 chronic exposure to O3 0.2 ppm and higher. The cilia were either completely absent or
18 had become noticeably shorter or blunt. After placing these animals in a clean-air
19 environment, the structurally damaged cilia regenerated and appeared normal (U.S. EPA.
20 1986). Based on such morphological observations, related effects such as ciliostasis,
21 increased mucus secretions, and a slowing of mucociliary transport rates might be
22 expected. However, no measurable changes in ciliary beating activity have been reported
23 due to O3 exposure alone. Essentially no data are available on the effects of prolonged
24 exposure to O3 on ciliary functional activity or on mucociliary transport rates measured in
25 the intact animal. In general, functional studies of mucociliary transport have observed a
26 delay in particle clearance soon after acute exposure. Decreased clearance is more
27 evident at higher doses (1 ppm), and there is some evidence of tolerance/adaptation for
28 these effects (U.S. EPA, 1986). However, no recent studies have evaluated the effects of
29 O3 on mucociliary clearance.
6.2.5.2 Alveolobronchiolar Transport Mechanism
30 In addition to the transport of particles deposited on the mucous surface layer of the
31 conducting airways, particles deposited in the deep lung may be removed either up the
32 respiratory tract or through interstitial pathways to the lymphatic system. The pivotal
33 mechanism of alveolobronchiolar transport involves the movement of AMs with
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1 phagocytized particles to the bottom of the mucociliary escalator. Failure of the AMs to
2 phagocytize and sequester the deposited particles from the vulnerable respiratory
3 membrane can lead to particle entry into the interstitial spaces. Once lodged in the
4 interstitium, particle removal is more difficult and, depending on the toxic or infectious
5 nature of the particle, its interstitial location may allow the particle to set up a focus for
6 pathologic processes. Although some studies show reduced early (tracheobronchial)
7 clearance after O3 exposure, late (alveolar) clearance of deposited material is accelerated,
8 presumably due to macrophage influx (which in itself can be damaging due to proteases
9 and oxidative reactions in these cells). In an important older study investigating the
10 effects of longer term O3 exposure on alveolobronchiolar clearance, rats were exposed to
11 an urban pattern of O3 (continuous 0.06 ppm, 7 days/week with a slow rise to a peak of
12 0.25 ppm and subsequent decrease to 0.06 ppm over a 9 h period for 5 days/week) for
13 6 weeks and were exposed 3 days later to chrysotile asbestos, which can cause pulmonary
14 fibrosis and neoplasia (Pinkerton et al. 1989). After 30 days, the lungs of the O3-exposed
15 animals had twice the number and mass of asbestos fibers as the air-exposed rats. New
16 evaluations of O3 effects on alveolar clearance have not been performed.
6.2.5.3 Alveolar Macrophages
17 Within the gaseous exchange region of the lung, the first line of defense against
18 microorganisms and nonviable particles that reach the alveolar surface is the AM. This
19 resident phagocyte is responsible for a variety of activities, including the detoxification
20 and removal of inhaled particles, maintenance of pulmonary sterility via destruction of
21 microorganisms, and interaction with lymphocytes for immunologic protection. Under
22 normal conditions, AMs seek out particles deposited on the alveolar surface and ingest
23 them, thereby sequestering the particles from the vulnerable respiratory membrane. To
24 adequately fulfill their defense function, the AMs must maintain active mobility, a high
25 degree of phagocytic activity, and an optimally functioning biochemical and enzyme
26 system for bactericidal activity and degradation of ingested material. As discussed in
27 previous AQCDs, short periods of O3 exposure can cause a reduction in the number of
28 free AMs available for pulmonary defense, and these AMs are more fragile, less
29 phagocytic, and have decreased lysosomal enzyme activities required for killing
30 pathogens. For example, in results from earlier work in rabbits, a 2-h exposure to 0.1 ppm
31 O3 inhibited phagocytosis and a 3-h exposure to 0.25 ppm decreased lysosomal enzyme
32 activities (Driscoll et al.. 1987; Hurst et al.. 1970). Similarly, AMs from rats exposed to
33 0.1 ppm O3 for 1 or 3 weeks exhibited reduced hydrogen peroxide production (Cohen et
34 al.. 2002). A controlled human exposure study reported decrements in the ability of
3 5 alveolar macrophages to phagocytize yeast following exposure of healthy volunteers to
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1 80 to 100 ppb O3 for 6.6-h during moderate exercise (Devlin et al., 1991). Although the
2 percentage of phagocytosis-capable macrophages was unchanged by O3 exposure, the
3 number of yeast engulfed was reduced when phagocytosis was complement-dependent.
4 However, there was no difference in the ability of macrophages to produce superoxide
5 anion after O3 exposure. These results are consistent with those from another controlled
6 human exposure study in which no changes in the level of lysosomal enzymes or
7 superoxide anion production were observed in macrophages lavaged from healthy human
8 subjects exposed to 400 ppb O3 for 2 h with heavy intermittent exercise (Koren et al..
9 1989). More recently, Lay et al. (2007) observed no difference in phagocytic activity or
10 oxidative burst capacity in macrophages or monocytes from sputum or blood collected
11 from healthy volunteers after a 2-hour exposure to 400 ppb O3 with moderate intermittent
12 exercise. However, another study found that oxidative burst and phagocytic activity in
13 macrophages increased in GSTM1 null subjects compared to GSTM1 positive subjects,
14 who had relatively unchanged macrophage function parameters after an O3 exposure
15 identical to that of Lay et al. described above (Alexis etal.. 2009). Collectively, these
16 studies demonstrate that O3 can affect multiple steps or aspects required for proper
17 macrophage function, but any concentration-response relationship appears complex and
18 genotype may be a consideration. A few other recent studies have evaluated ozone's
19 effects on macrophage function, but these are of questionable relevance due to the use of
20 in vitro exposure systems and amphibian animal models (Mikerov et al., 2008b; Dohm et
21 al.. 2005: Klestadt et al.. 2005).
6.2.5.4 Infection and Adaptive Immunity
General Effects on the Immune System
22 The effects of O3 on the immune system are complex and dependent on the exposure
23 regimen and the observation period. According to toxicological studies it appears that the
24 T-cell-dependent functions of the immune system are more affected than B-cell-
25 dependent functions (U.S. EPA. 2006b). Generally, there is an early immunosuppressive
26 effect that subsides with continued O3 exposure, resulting in either a return to normal
27 responses or an enhancement of immune responses. However, this is not always the case
28 as Aranyi (1983) showed decreased T-cell mitogen reactions in mice after subchronic
29 (90-day) exposure to 0.1 ppm O3. Earlier studies report changes in cell populations in
30 lymphatic tissues (U.S. EPA. 2006b). A more recent study in mice demonstrated that
31 numbers of certain T cell subsets in the spleen were reduced after exposure to 0.6 ppm O3
32 (lOh/day x 15d) (Feng et al.. 2006).
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1 The inflammatory effects of O3 involve the innate immune system, and as such can affect
2 adaptive (or acquired) immunity via alterations in antigen presentation and costimulation
3 by innate immune cells such as macrophages and dendritic cells. Several recent
4 controlled human exposure studies demonstrate increased expression of molecules
5 involved in antigen presentation or costimulation. Lay et al. (2007) collected sputum
6 monocytes from healthy volunteers exposed to 400 ppb O3 for 2 h with moderate
7 intermittent exercise and detected increases in HLA-DR, used to present antigen to T
8 cells, and CD86, a costimulatory marker necessary for T cell activation. Upregulation of
9 HLA-DR was also observed by Alexis et al. (2009) in sputum dendritic cells and
10 macrophages from GSTM1 null subjects exposed to 400 ppb O3 for 2 h with moderate
11 intermittent exercise. On airway monocytes from healthy volunteers 24 hours after
12 exposure to 80 ppb O3 for 6.6 h with moderate intermittent exercise, HLA-DR, CD86,
13 and CD 14 (a molecule involved in bacterial endotoxin reacitivity) were increased,
14 whereas CD80, a costimulatory molecule of more heterogeneous function, was decreased
15 (Alexis et al.. 2010). Patterns of expression on macrophages were similar, except that
16 HLA-DR was found to be significantly decreased after O3 exposure and CD86 was not
17 significantly altered. An increase in IL-12p70, a macrophage and dendritic cell product
18 that activates T cells, was correlated with increased numbers of dendritic cells. It should
19 be noted that these results are reported as comparisons to baseline as there was no clean
20 air control (Alexis et al.. 2010; Alexis et al.. 2009). Another controlled human exposure
21 study reported no increase in IL-12p70 in sputum from healthy, atopic, or atopic
22 asthmatic subjects following a 2-hour exposure to 400 ppb O3 with intermittent moderate
23 exercise (Hernandez etal.. 2010). Levels of HLA-DR, CD14 and CD86 were not
24 increased on macrophages collected from any of these subjects. It is difficult to compare
25 these results to those of Lay et al. (2007) and Alexis et al. (2010) due to differences in O3
26 concentration, cell type examined, and timing of postexposure analysis.
27 Although no controlled human exposure studies have examined the effects of O3 on the
28 ability to mount antigen-specific responses, upregulation of markers associated with
29 innate immune activation and antigen presentation could potentially enhance adaptive
30 immunity and increase immunologic responses to antigen. While this may bolster
31 defenses against infection, it also may enhance allergic responses (Section 6.2.6).
32 In animal models, O3 has been found to alter responses to antigenic stimulation. For
33 example, antibody responses to a T-cell-dependent antigen were suppressed after a
34 56-day exposure of mice to 0.8 ppm O3, and a 14-day exposure to 0.5 ppm O3 decreased
35 the antiviral antibody response following influenza virus infection (Jakab and Hmieleski.
36 1988); the latter impairment may pave the way for lowered resistance to reinfection. The
37 immune response is highly influenced by the temporal relationship between O3 exposure
38 and antigenic stimulation. When O3 exposure preceded Listeria infection, there were no
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1 effects on delayed-type hypersensitivity or splenic lymphoproliferative responses;
2 however, when O3 exposure occurred during or after Listeria infection was initiated,
3 these immune responses were suppressed (Van Loveren et al., 1988). In another study, a
4 reduction in mitogen activated T-cell proliferation was observed after exposure to
5 0.6 ppm for 15 days, and could be ameliorated by antioxidant supplementation. Antigen-
6 specific proliferation decreased by 60%, indicating attenuation of the acquired immunity
7 needed for subsequent memory responses (Feng etal.. 2006). O3 exposure also skewed
8 the ex-vivo cytokine responses elicited by non-specific stimulation toward inflammation,
9 decreasing IL-2 and increasing IFN-y. Modest decreases in immune function assessed in
10 the offspring of O3-exposed dams (mice) were observed by Sharkhuu et al. (2011). The
11 ability to mount delayed-type hypersensitivity responses was significantly suppressed in
12 42 day-old offspring when dams were exposed to 0.8 or 1.2 ppm O3, but not 0.4 ppm,
13 from gestational day 9-18. Humoral responses to immunization with sheep red blood
14 cells were unaffected, as were other immune parameters such as splenic populations of
15 CD45+ T cells, iNKT cells, and levels of IFN-y, IL-4, and IL-17 in the BALF. Generally,
16 continuous exposure to O3 impairs immune responses for the first several days of
17 exposure, followed by an adaptation to O3 that allows a return of normal immune
18 responses. Most species show little effect of O3 exposures prior to immunization, but
19 show a suppression of responses to antigen in O3 exposures post-immunization.
Microbial Infection
Bacterial infection
20 A relatively large body of evidence shows that O3 increases susceptibility to bacterial
21 infections. The majority of studies in this area were conducted before the 1996 O3 AQCD
22 was published and many are included in Table 6-9 on p. 6-53 of that document. Known
23 contributing factors are impaired mucociliary streaming, altered chemotaxis/motility,
24 defective phagocytosis of bacteria, decreased production of lysosomal enzymes or
25 superoxide radicals by alveolar macrophages, and decreased IFN^y levels. In animal
26 models of bacterial infection, exposure to 0.08 ppm O3 increases streptococcus-induced
27 mortality, regardless of whether O3 exposure precedes or follows infection (Miller et al..
28 1978; Coffin and Gardner. 1972; Coffin et al., 1967). Increases in mortality are due to the
29 infectious agent, thereby reflecting functional impairment of host defenses. Exercise and
30 copollutants can enhance ozone's effects in infectivity models. Although both mice and
31 rats exhibit impaired bactericidal macrophage activity after O3 exposure, mortality due to
32 infection is only observed in mice. Additionally, although mice and humans share many
33 host defense mechanisms, there is little compelling evidence from epidemiologic studies
34 (Section 6.2.7.3).
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Viral infection
1 Only a few studies, described in previous AQCDs, have examined the effects of O3
2 exposure on the outcome of viral respiratory infection (see Table 6-9 on p. 6-53 of the
3 1996 O3 AQCD. Some studies show increased mortality, while others show diminished
4 severity and increased survival time. There is little to no evidence from studies of animals
5 or humans to suggest that O3 increases the incidence of respiratory viral infection in
6 humans. In human volunteers infected with rhinovirus prior to O3 exposure (0.3 ppm for
7 5 consecutive days), no effect on viral titers, IFN-y production, or blood lymphocyte
8 proliferative responses to viral antigen was observed (Henderson et al.. 1988). In vitro
9 cell culture studies of human bronchial epithelial cells indicate O3-induced exacerbation
10 of human rhinovirus infection (Spannhake et al.. 2002). but this is of limited relevance.
11 Newer studies on the interactions of O3 and viral infections have not been published.
12 Natural killer (NK) cells, which destroy virally infected cells and tumors in the lung,
13 appear to be inhibited by higher concentrations of O3 and either unaffected or stimulated
14 at lower concentrations. Several studies show decreases in NK cell activity following
15 acute exposures ranging from 0.8 to 1 ppm (Gilmour and Jakab. 1991; Van Loveren et
16 al.. 1990; Burleson et al.. 1989). However, Van Loveren et al. (1990) showed that a
17 1-week exposure to 0.2 or 0.4 ppm O3 increased NK cell activity, and an urban pattern of
18 exposure (base of 0.06 ppm with peaks of 0.25 ppm) had no effect on NK cell activity
19 after 1, 3, 13, 52, or 78 weeks of exposure (Selgrade et al.. 1990). A more recent study
20 demonstrated a 35% reduction in NK cell activity after exposure of mice to 0.6 ppm O3
21 (lOh/day x 15d) (Feng et al.. 2006). The defective IL-2 production demonstrated in this
22 study may impair NK cell activation. Alternatively, NK cell surface charge may be
23 altered by ROS, decreasing their adherence to target cells (Nakamura and Matsunaga.
24 1998).
Summary: Infections
25 Taken as a whole, the data clearly indicate that an acute O3 exposure impairs the host
26 defense capability of both humans and animals, primarily by depressing alveolar
27 macrophage function and perhaps also by decreasing mucociliary clearance of inhaled
28 particles and microorganisms. This suggests that humans exposed to O3 could be
29 predisposed to bacterial infections in the lower respiratory tract. The seriousness of such
30 infections may depend on how quickly bacteria develop virulence factors and how
31 rapidly PMNs are mobilized to compensate for the deficit in alveolar macrophage
32 function. To date, a limited number of epidemiologic studies have examined associations
33 between O3 exposure and HA/ED for respiratory infection, pneumonia, or influenza.
34 Results have been mixed, and in some cases conflicting (see Sections 6.2.7.2 and
35 6.2.7.3). With the exception of influenza, it is difficult to ascertain whether cases of
36 respiratory infection or pneumonia are of viral or bacterial etiology. A study that
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1 examined the association between O3 exposure and respiratory hospital admissions in
2 response to an increase in influenza intensity did observe an increase in respiratory
3 hospital admissions (Wong et al.. 2009). but information from toxicological studies of O3
4 and viral infections is ambiguous.
6.2.6 Allergic and Asthma-Related Responses
5 Effects resulting from combined exposures to O3 and allergens have been studied in a
6 variety of animal species, generally as models of experimental asthma. Pulmonary
7 function and airways hyperresponsiveness in animal models of asthma are discussed in
8 Sections 6.2.1.7 and 6.2.2.2. Previous evidence indicates that O3 exposure skews immune
9 responses toward an allergic phenotype. For example, Gershwin et al. (1981) reported
10 that O3 (0.8 and 0.5 ppm for 4 days) exposure caused a 34-fold increase in the number of
11 IgE (allergic antibody)-containing cells in the lungs of mice. In general, the number of
12 IgE-containing cells correlated positively with levels of anaphylactic sensitivity. In
13 humans, allergic rhinoconjunctivitis symptoms are associated with increases in ambient
14 O3 concentrations (Riediker et al., 2001). Recent controlled human exposure studies have
15 observed O3-induced changes indicating allergic skewing. Airway eosinophils, which
16 participate in allergic disease and inflammation, were observed to increase in atopic,
17 mildly asthmatic volunteers 18 h following a 7.6-hour exposure to 160 ppb O3 with light
18 intermittent exercise (Peden et al., 1997). No increase in airway eosinophils was observed
19 4 h after exposure of healthy, atopic, or atopic asthmatic subjects to 400 ppb O3 for 2 h
20 with moderate intermittent exercise (Hernandez et al.. 2010). However, atopic subjects
21 did exhibit increased IL-5, a cytokine involved in eosinophil recruitment and activation,
22 suggesting that perhaps these two studies observed the same effect at different time
23 points. Several epidemiologic studies discussed in Section 7.2.5 describe an association
24 between eosinophils and long-term O3 exposure, consistent with chronic exposure studies
25 in non-human primates. Hernandez et al. (2010) also observed increased expression of
26 high and low affinity IgE receptors on sputum macrophages from atopic asthmatics,
27 which may enhance IgE-dependent inflammation. Sputum levels of IL-4 and IL-13, both
28 pro-allergic cytokines that aid in the production of IgE,were unaltered in any group. The
29 lack of increase in IL-4 levels in sputum reported by Hernandez et al., along with
30 increased IL-5, is consistent with results from Bosson et al. (2003). in which IL-5 (but not
31 IL-4 levels) increased in bronchial epithelial biopsy specimens following exposure of
32 mild atopic asthmatics to 200 ppb O3 for 2 h with moderate intermittent exercise. IL-5
33 was not elevated in specimens obtained from healthy (non-asthmatic) O3-exposed
34 subjects. Collectively, findings from these studies suggest that O3 can induce or enhance
35 certain components of allergic inflammation in atopic and atopic asthmatic individuals.
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1 Ozone enhances inflammatory and allergic responses to allergen challenge in sensitized
2 animals. Short-term exposure (2 days) to 1 ppm O3 exacerbated allergic rhinitis and lower
3 airway allergic inflammation in Brown Norway rats, a rat strain that is comparatively less
4 sensitive to O3 than other rats or humans (Wagner et al.. 2009; Wagner et al.. 2007).
5 OVA-sensitized rats were intranasally challenged with OVA on days 1 and 2, and
6 exposed to 0 or 1 ppm O3 (8 h/day) on days 4 and 5. Analysis at day 6 indicated that O3
7 exposure enhanced intraepithelial mucosubstances in the nose and airways, induced cys-
8 LTs, MCP-1, and IL-6 production in BALF, and upregulated expression of the
9 proallergic cytokines IL-5 and IL-13. These changes were not evident in non-allergic
10 controls. All of these responses were blunted by gamma-tocopherol (yT; vitamin E)
11 therapy. yT neutralizes oxidized lipid radicals, and protects lipids and proteins from
12 nitrosative damage from NO-derived metabolites. Farraj et al. (2010) exposed allergen-
13 sensitized adult male BALB/c mice to 0.5 ppm O3 for 5 hours once per week for 4 weeks.
14 Ozone exposure and O3/DEP (2.0 mg/m3) co-exposure of OVA-sensitized mice elicited
15 significantly greater serum IgE levels than in DEP-exposed OVA-sensitized mice (98%
16 and 89% increases, respectively). Ozone slightly enhanced levels of BAL IL-5, but
17 despite increases in IgE, caused a significant decrease in BAL IL-4 levels. IL-10, IL-13,
18 and IFN-y levels were unaffected. Lung resistance and elastance were unaffected in
19 allergen sensitized mice exposed solely to 0.5 ppm O3 once a week for 4 weeks (Farraj et
20 al.. 2010). However, co-exposure to O3 and diesel exhaust particles increased lung
21 resistance.
22 In addition to exacerbating existing allergic responses, O3 can also act as an adjuvant to
23 produce sensitization in the respiratory tract. In a model of murine asthma, using OVA
24 free of detectable endotoxin, inclusion of 1 ppm O3 during the initial exposures to OVA
25 (2 h, days 1 and 6) enhanced the inflammatory and allergic responses to subsequent
26 allergen challenge (Hollingsworth et al.. 2010). Compared to air exposed animals, O3
27 exposed mice exhibited significantly higher levels of total cells, macrophages,
28 eosinophils, and PMNs in BALF, and increased total serum IgE. Pro-allergic cytokines
29 IL-4, and IL-5 were also significantly elevated, along with pleiotropic Th2 cytokine IL-9
30 (associated with bronchial hyperresponsiveness) and pro-inflammatory IL-17, produced
31 by activated T cells. Based on lower inflammatory, IgE, and cytokine responses in Toll-
32 like receptor 4 deficient mice, the effects of O3 seem to be dependent on TLR 4 signaling,
33 as are a number of other biological responses to O3 according to studies by Hollingsworth
34 et al. (2004). Kleeberger et al.QOOO) and Garanziotis et al. (2010). The involvement of
35 TLR 4, along with its endogenous ligand, hyaluronan, in O3-induced responses described
36 in these studies has been corroborated by a controlled human exposure study by
37 Hernandez et al. (2010). who found increased TLR 4 expression and elevated levels of
38 hyaluronic acid in atopic and atopic asthmatic volunteers exposed to 400 ppb O3. This
39 pathway is discussed in more detail in Chapter 5. Examination of dendritic cells (DCs)
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1 from the draining thoracic lymph nodes indicated that O3 did not enhance the migration
2 of DCs from the lungs to the lymph nodes, nor did it alter the expression of functional
3 DC markers such as CD40, MHC class II, or CD83. However, O3 did increase expression
4 of CD86, which is generally associated with Th2 responses and is detected at higher
5 levels on DCs from allergic asthmatics compared to those from healthy donors (Chen et
6 al.. 20061)). Increased CD86 has also been observed on airway cells collected from
7 human subjects following exposure to O3 in studies by Lay et al. (2007) and Alexis et al.
8 (2009). but not Hernandez et al. (2010) (study details described in Section 6.2.5.4).
9 Ozone exposure during gestation has modest effects on allergy and asthma related
10 endpoints in adult offspring. When dams were exposed to 1.2 ppm O3 (but not 0.8 ppm)
11 from gestational day 9-18, some allergic and inflammatory responses to OVA
12 sensitization and challenge were reduced compared to air exposed controls. This included
13 IgE levels and eosinophils, and was only true of mice that were immunized early in life
14 (PND 3) as opposed to later (PND 42), perhaps due to the proximity of O3 and antigen
15 exposure. The effects of gestational O3 exposure on immune function have not been
16 widely studied, and although reductions in allergic endpoints are not generally observed
17 in association with O3, other parameters of immune function were found to be reduced, so
18 a more global immunosuppression may underlie these effects.
19 In addition to ozone's pro-allergic effects, it could also make airborne allergens more
20 allergenic. When combined with NO2, O3 has been shown to enhance nitration of
21 common protein allergens, which may increase their allergenicity (Franze et al.. 2005).
6.2.7 Hospital Admissions, Emergency Department Visits, and Physicians
Visits
6.2.7.1 Summary of Findings from 2006 Ozone AQCD
22 The 2006 O3 AQCD evaluated numerous respiratory ED visits and hospital admissions
23 studies, which consisted primarily of time-series studies conducted in the U.S., Canada,
24 Europe, South America, Australia and Asia. Upon collectively evaluating the scientific
25 evidence, the 2006 O3 AQCD concluded that "the overall evidence supports a causal
26 relationship between acute ambient O3 exposures and increased respiratory morbidity
27 resulting in increased ED visits and [hospital admissions] during the warm season" (U.S.
28 EPA. 2006b). This conclusion is "strongly supported by the human clinical, animal
29 toxicologicfal], and epidemiologic evidence for [O3-induced] lung function decrements,
30 increased respiratory symptoms, airway inflammation, and airway hyperreactivity" (U.S.
31 EPA. 2006b).
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1 Since the completion of the 2006 O3 AQCD, relatively fewer studies conducted in the
2 U.S., Canada, and Europe have examined the association between short-term exposure to
3 ambient O3 and respiratory hospital admissions and ED visits with a growing number of
4 studies having been conducted in Asia. This section focuses primarily on multicity
5 studies because they examine the effect of O3 on respiratory-related hospital admissions
6 and ED visits over a large geographic area using a consistent statistical methodology.
7 Single-city studies that encompass a large number of hospital admissions or ED visits, or
8 included a long study-duration were also evaluated because these studies have more
9 power to detect whether an association exists between short-term O3 exposure and
10 respiratory hospital admissions and ED visits compared to smaller single-city studies.
11 Additional single-city studies were also evaluated within this section, if they were
12 conducted in locations not represented by the larger single-city and multicity studies, or
13 examined population-specific characteristics not included in the larger studies that may
14 modify the association between short-term O3 exposure and respiratory-related hospital
15 admissions or ED visits. The remaining single-city studies identified were not evaluated
16 in this section due to factors such as inadequate study design or insufficient sample size.
17 It should be mentioned that when examining the association between short-term O3
18 exposure and respiratory health effects that require medical attention, it is important to
19 distinguish between hospital admissions and ED visits. This is because it is likely that a
20 small percentage of respiratory ED visits will be admitted to the hospital; therefore,
21 respiratory ED visits may represent potentially less serious, but more common outcomes.
22 As a result, in the following sections respiratory hospital admission and ED visit studies
23 are evaluated individually. Additionally, within each section, results are presented as
24 either a collection of respiratory diagnoses or as individual diseases (e.g., asthma, COPD,
25 pneumonia and other respiratory infections) in order to evaluate the potential effect of
26 short-term O3 exposure on each respiratory-related outcome. The ICD codes (i.e., ICD-9
27 or ICD-10) that encompass each of these endpoints are presented in Table 6-25 along
28 with the air quality characteristics of the city, or across all cities, included in each study
29 evaluated in this section.
Draft - Do Not Cite or Quote 6-111 September 2011
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Table 6-25 Mean and upper percentile concentrations of respiratory-related
hospital admission and emergency department visit studies
evaluated
Study
Katsouyanni et
al. (2009)"'°
Cakmaketal.
(2006b)
Biggeri et al.
/9nn^°
\£.\J\J\Jf
Dales etal.
(2006)
Lin et al. (2008a)
Wong etal.
(2009)°
Medina-Ramon
etal. (2006)h
Yang etal.
(2005b)
Zanobetti and
Schwartz
(2006)"
Silverman and
lto(2010)b
Stieb et al.
(2009)
Tolbertetal.
(2007)
Darrow et al.
(201 1b)
Villeneuveetal.
(2007)b
Ito et al. (2007b)
Location
90 U.S. cities
(NMMAPS)d
32 European
cities (APHEA)d
12 Canadian
cities
10 Canadian
cities
4 Italian cities'
11 Canadian
cities
11 New York
regions
Hong Kong
36 U.S. cities
Vancouver,
Canada
Boston, MA
New York, NY
7 Canadian
cities
Atlanta, GA
Atlanta, GA
Alberta, CAN
New York, NY
Type of Visit (ICD9/10)
Hospital Admissions:
NMMAPS: All respiratory (460-
519)
APHEA:AII respiratory (460-519)
12 Canadian cities: All
respiratory (460-51 9)e
Hospital Admissions:
All respiratory (466, 480-486,
490,491,492,493,494,496)
Hospital Admissions:
All respiratory (460-51 9)
Hospital Admissions:
Respiratory disorders (486,
768.9, 769, 770.8, 786, 799.0,
799.1)
Hospital Admissions:
Respiratory diseases (466, 490-
493, 496)
Hospital Admissions:
All respiratory (460-519)
Hospital Admissions:
COPD (490-496, excluding 493)
Pneumonia (480-487)
Hospital Admissions:
COPD (490-492, 494, 496)
Hospital Admissions:
Pneumonia (480-487)
Hospital Admissions:
Asthma (493)
Emergency Department Visits:
Asthma (493)
COPD (490-492, 494-496)
Respiratory infection (464, 466,
480-487)
Emergency Department Visits:
All respiratory (460-465, 460.0,
466.1,466.11,466.19,477,480-
486,491,492,493,496,786.07,
786.09)
Emergency Department Visits:
All respiratory (460-466, 477,
480-486,491,492,493,496,
786.09)
Emergency Department Visits:
Asthma (493)
Emergency Department Visits:
Asthma (493)
Averaging
Time
1-hmax
24-h avg
8-h max
24-h avg
8-h maxs
8-h maxs
8-h max
24-h avg
24-h avg
8-h max
24-h avg
8-h max
8-h max
1-hmax
24-h avg
Commute
Day-time
Night-time
8-h max
8-h max
Mean
Concentration (ppb)a
NMMAPS:
50th: 34.9-60.0
APHEA:
50th: 11. 0-38.1
12 Canadian cities:
50th: 6.7-8.3
17.4
Warm season
(May-September): 5.7-60.0
17.0
44.1
18.8
Warm
(May-September): 45.8
Cool
(October-April): 27.6
All year: 14.1
Winter
(January-March): 13.2
Spring
(April-June): 19.4
Summer
(July-September): 13.8
Fall
(October-December): 10.0
22.4
Warm
(April-August):41.0
18.4
Warm: 53.0
Warm
(March-October):
8-h max: 53
1-hmax: 62
24-h avg: 30
Commute: 35'
Day-time: 45' .
Night-time: 14'
Summer
(April-September): 38.0
Winter
(October-March): 24.3
All year: 30.4
Warm
(April-September): 42.7
Cold
(October-March): 18.0
Upper Percentile
Concentrations (ppb)a
NMMAPS:
75th: 46.8-68.8
APHEA:
75th: 15.3-49.4
12 Canadian cities:
75th: 8.9-12.4
Max: 38.0-79.0
95th: 86. 1-90.0
Max: 107.5-115.1
95th: 24.9-46.0
75th: 54.0
Max: 21 7.0
75th: 25.9
Max: 100.3
NR
Max: 38.6
75th: 31.0
95th: 47.6
75th: 53
90th: 68
75th: 19.3-28.6
75th: 67.0
90th: 82.1
Max: 147.5
8-h 24-h avg:
max: 75th: 37
75th: 67 Max: 81
Max: Commute:
148 75th: 45
1-h Max: 106
max:
75th: 76
Max:
180
Summer:
75th: 46.0
Winter:
75th: 31 .5
All year:
95th: 68.0
Warm months:
95th: 77.0
Cold months:
95th: 33.0
Day-
time:
75th: 58
Max:
123
Night-
time:
75th: 22
Max: 64
Draft - Do Not Cite or Quote
6-112
September 2011
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Study
Strickland etal.
(2010)
Mar and Koenig
etal. (2009)
Arbexetal.
(2009)
Orazzoetal.
(2009)°
Burra et al.
(2009)
Villeneuve etal.
(2006b)
Sinclair etal.
(2010)'
Location
Atlanta, GA
Seattle, WA
Sao Paulo,
Brazil
6 Italian cities
Toronto,
Canada
Toronto,
Canada
Atlanta, GA
Type of Visit (ICD9/10)
Emergency Department Visits:
Asthma (493)
Wheeze (786.07 after 10/1/98,
786.09 before 10/1/98)
Emergency Department Visits:
Asthma (493-493.9)
Emergency Department Visits:
COPD (J40-44)
Emergency Department Visits:
Wheezing
Physician Visits:
Asthma (493)
Physician Visits:
Allergic rhinitis (177)
Physician Visits:
Asthma
Upper respiratory infection
Lower respiratory infection
Averaging
Time
8-h max
1-h max
8-h max
1-h max
8-h max"
1-h max
8-h max
8-h max
Mean
Concentration (ppb)a
All year: 45.41
Warm
(May-October): 55.2J
Cold
(November-April): 34.5J
Warm (May-October):
1-h max: 38.6
8-h max: 32.2
48.8
Summer
(April-September):
21.1-44.3
Winter
(October-March): 11.5-27.9
33.3
30.0
Total Study Period:
All-year: 44.0
25 mo Period:
All-year: 47.9
Warm: 61 .2
Cold: 27.8
28 mo Period:
All-year: 40.7
Warm: 51 .8
Cold: 26.0
Upper Percentile
Concentrations (ppb)a
NR
75th:
1-h max: 45.5
8-h max: 39.2
75th: 61.0
Max: 143.8
NR
95th: 66
Max: 121
Max: 98.7
NR
aSome studies did not present an overall value for the mean, middle and/or upper percentiles of the 03 distribution; as a result, the range of the
mean, middle, and/or upper percentiles across all of the cities included in the study are presented.
bStudy only presented median concentrations.
°Study presented concentrations as ug/m3 Concentration was converted to ppb using the conversion factor of 0.51 assuming standard
temperature (25°C) and pressure (1 atm).
dA subset of the European and U.S. cities included in the mortality analyses were used in the hospital admissions analyses: 8 of the 32 European
cities and 14 of 90 U.S. cities.
eHospital admission data was coded using three classifications (ICD-10-CA, ICD-9, and ICD-9-CM). Attempts were made by the original
investigators to convert diagnosis from ICD-10-CA back to ICD-9.
'Only 4 of the 8 cities included in the study collected 03 data.
903 measured from 10:00 a.m. to 6:00 p.m.
hOnly 35 of the 36 cities included in the analysis had 03 data.
'Commute (7:00 a.m. to 10:00 a.m., 4:00 p.m. to 7:00 p.m.); Day-time (8:00 a.m. to 7:00 p.m.); Night-time (12:00 a.m. to 6:00 a.m.).
'Means represent population-weighted 03 concentrations.
k03 measured from 8:00 a.m. to 4:00 p.m.
'This study did not report the ICD codes used for the conditions examined. The 25-month period represents August 1998-August 2000, and the
28-month period represents September 2000-December 2002. This study defined the warm months as April - October and the cold months as
November-March.
6.2.7.2 Hospital Admission Studies
1
2
3
4
5
Respiratory Diseases
The association between exposure to an air pollutant, such as O3, and daily respiratory-
related hospital admissions has primarily been examined using all respiratory-related
hospital admissions within the range of ICD-9 codes 460-519. Newly identified studies
attempt to further examine the effect of O3 exposure on respiratory-related hospital
admissions through a multicity design that examines O3 effects across countries using a
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1 standardized methodology; multicity studies that examine effects within one country; and
2 multi- and single-city studies that attempt to examine potential modifiers of the O3-
3 respiratory-related hospital admission relationship.
4 The Air Pollution and Health: A European and North American Approach (APHENA)
5 study combined data from existing multicity study databases from Canada, Europe
6 (APHEA2) (Katsouvanni et al.. 2001). and the U.S. (NMMAPS) (Samet et al.. 2000) in
7 order to "develop more reliable estimates of the potential acute effects of air pollution on
8 human health [and] provide a common basis for [the] comparison of risks across
9 geographic areas" (Katsouyanni et al.. 2009). In an attempt to address both of these
10 issues, the investigators conducted extensive sensitivity analyses to evaluate the
11 robustness of the results to different model specifications (e.g., penalized splines [PS]
12 versus natural splines [NS]) and the extent of smoothing to control for seasonal and
13 temporal trends. The trend analyses consisted of subjecting the models to varying extent
14 of smoothing selected either a priori (e.g., 3 df/year, 8 df/year, and 12 df/year) or by
15 using the absolute sum of the residuals of the partial autocorrelation function (PACF).
16 However, the investigators did not identify the model they deemed to be the most
17 appropriate for comparing the results across study locations. As a result, when discussing
18 the results across the three study locations below, the 8 df/year results are presented for
19 both the PS and NS models because: (1) 8 df/year is most consistent with the extent of
20 temporal adjustment used in previous and recent large multicity studies in the U.S. (e.g.,
21 NMMAPS); (2) the risk estimates for 8 df/year and 12 df/year are comparable for all
22 three locations; (3) the models that used the PACF method did not report the actual
23 degrees of freedom chosen; and (4) the 3 df/year and the PACF method resulted in
24 negative O3 risk estimates, which is inconsistent with the results obtained using more
25 aggressive seasonal adjustments. Additionally, when comparing results across studies in
26 figures, only the results from one of the spline models (e.g., NS) are presented because it
27 has been previously demonstrated that alternative spline models result in relatively
28 similar effect estimates (HEI. 2003). However, it should be noted that the underlying data
29 and model specifications could result in varying degrees of bias and precision in effect
30 estimates with different spline models (Ostro et al.. 2006).
31 Katsouyanni et al. (2009) examined respiratory hospital admissions for people aged
32 65 years and older using 1-h max O3 data. The extent of hospital admission and O3 data
33 varied across the 3 datasets: Canadian dataset included 12 cities with data for 3 years
34 (1993-1996) per city; European dataset included 8 cities with each city having data for
35 between 2 and 8 years from 1988-1997; and U.S. dataset included 14 cities with each city
36 having data for between 4 and 10 years from 1985-1994 and 7 cities having only summer
37 O3 data. The investigators used a three-stage hierarchical model to account for within-
3 8 city, within region, and between region variability. Results were presented individually
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1 for each region (Figure 6-14; Table 6-26). Ozone and PM10 concentrations were weakly
2 correlated in all locations in the summer (r=0.27-0.40), but not in the winter.
3 In the Canadian cities, using all-year data, a 40 ppb increase in 1-h max O3
4 concentrations at lag 0-1 was associated with an increase in respiratory hospital
5 admissions of 8.9% (95% CI: 0.79, 16.8%) in a PS model and 8.1% (95% CI: 0.24,
6 16.8%) in a NS model (Katsouyanni et al., 2009). The results were somewhat sensitive to
7 the lag day selected, reduced when using a single-day lag (e.g., lag 1) (PS: 6.0%; NS:
8 5.5%) and increased when using a distributed lag model (PS: 18.6%; NS: 20.4%). When
9 adjusting for PMi0, the magnitude of the effect estimate was slightly larger in the NS
10 model (5.1% [95% CI: -6.6, 18.6%]) compared to the PS model (3.1% [95% CI: -8.3,
11 15.9%]); however, the copollutant analysis was only conducted using a 1-day lag. The
12 large confidence intervals for both models could be attributed to the reduction in days
13 included in the copollutant analyses as a result of the every-6th-day PM sampling
14 schedule. When restricting the analysis to the summer months, stronger associations were
15 observed between O3 and respiratory hospital admissions across the lags examined,
16 ranging from -22 to 37% (the study does not specify whether these effect estimates are
17 from a NS or PS model). Because O3 concentrations across the cities included in the
18 Canadian dataset (Katsouyanni et al. (2009) are low (median concentrations ranging from
19 6.7-8.3 ppb [Table 6-25]), the standardized increment of 40 ppb for a 1-h max increase in
20 O3 concentrations does not accurately reflect the observed risk of O3-related respiratory
21 hospital admissions. Although this increment adequately characterizes the distribution of
22 1-h max O3 concentrations across the U.S. and European datasets, it misrepresents the
23 observed O3 concentrations in the Canadian dataset. As a result in summary figures, for
24 comparability, effect estimates from the Canadian dataset are presented for both a 5.1 ppb
25 increase in 1-h max O3 concentrations (i.e., an approximate interquartile range [IQR]
26 increase in O3 concentrations across the Canadian cities) as well as the standardized
27 increment used throughout the ISA.
28 In Europe, weaker but positive associations were also observed in year round analyses;
29 2.9% (95% CI: 0.63, 5.0%) in the PS model and 1.6% (95% CI: -1.7, 4.2%) in the NS
30 model at lag 0-1 for a 40 ppb increase in 1-h max O3 concentrations (Katsouyanni et al..
31 2009). Additionally, at lag 1, associations between O3 and respiratory hospital admissions
32 were also reduced, but in contrast to the lag 0-1 analysis, greater effects were observed in
33 the NS model (2.9% [95% CI: 1.0, 4.9%]) compared to the PS model (1.5% [95% CI: -
34 2.2, 5.4]). Unlike the Canadian analysis, a distributed lag model provided limited
35 evidence of an association between O3 and respiratory hospital admissions. To compare
36 with the Canadian results, when adjusting for PM10 at lag 1, effect estimates were
37 increased in the PS model (2.5% [95% CI: 0.39-4.8%]) and remained robust in the NS
38 model (2.4% [95% CI: 0.08, 4.6%]). However, the European analysis also examined the
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1
2
3
4
5
6
7
effect of adjusting for PM10 at lag 0-1 and found results were attenuated in both models
(PS: 0.8% [95% CI: -2.3, 4.0%]; NS: 0.8% [95% CI: -1.8, 3.6%]). Unlike the Canadian
and U.S. datasets, the European dataset consisted of daily PM data. The investigators did
not observe stronger associations in the summer-only analyses for the European cities at
lag 0-1 (PS: 0.4% [95% CI: -3.2, 4.0%]; NS: 0.2% [95% CI: -3.3, 3.9%]), but did observe
some evidence for larger effects during the summer, an -2.5% increase, at lag 1 in both
models (the study does not present the extent of temporal smoothing used for these
models).
Location Lag
U.S. 1
1
0-1
0-1
DL(0-2)
0-1
1
Canada 1
la
1
la
0-1
0-la
DL(0-2)
DL(0-2)a
1
la
0-1
0-la
DL(0-2)
DL(0-2)a
Europe 1
1
0-1
0-1
DL(0-2)
1
0-1
% Increase
Black circles = all-year results; open circles = all-year results in copollutant model with PM10; and red circles = summer only
results. For Canada, lag days with an "a" next to them represent the risk estimates standardized to an approximate IQR of 5.1 ppb
for a 1-h max increase in ozone concentrations.
Figure 6-14 Percent increase in respiratory hospital admissions from natural
spline models for a 40 ppb increase in 1-h max ozone
concentrations for each location of the APHENA study.
10 -5 C
— •— All-Year
-0
— • Summer
• All Ynir
*-
Q
o-
• k
• ^
• ^
— •— Ail-Year
-• —
O
t
) 5 10 15 20 25 30 35 40
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Table 6-26 Corresponding effect estimates for Figure 6-14
Location Season Lag3
Copollutant % Increase (95% Cl)°
U.S.
All-year 1
1
0-1
0-1
DL(0-2)
Summer 0-1
1
Canada All-year 1
1a
1
1a
0-1
0-1 a
DL(0-2)
DL(0-2)a
Summer 1
1a
0-1
0-1 a
DL(0-2)
DL(0-2)a
Europe All-year 1
1
0-1
0-1
DL(0-2)
Summer 1
0-1
2.62 (0.63, 4.64)
PM10 2. 14 (-0.08, 4.40)
2.38 (0.00, 4.89)
PM10 1.42 (-1.33, 4.23)
3.34 (0.02-6.78)
2. 14 (-0.63, 4.97)
2.78 (-0.02, 5.71)
5.54 (-0.94, 12.4)
0.69 (-0.12, 1.50)a
PM10 5.13 (-6.62, 18.6)
PM10 0.64 (-0.87, 2.20)a
8.12(0.24, 16.8)
1 .00 (0.03, 2.00)a
20.4 (4.07, 40.2)
2.4(0.51,4.40)3
21.4(15.0,29.0)
2.50(1.80,3.30)3
32.0(18.6,47.7)
3.60(2.20,5.10)3
37.1 (11.5,67.5)
4.1 (1.40,6.80)3
2.94(1.02,4.89)
PM10 2.38 (0.08, 4.64)
1.58 (-1.71, 4.15)
PM10 0.87 (-1.79, 3.58)
0.79 (-4.46, 6.37)
2.46 (-0.63, 5.54)
0.24 (-3.32, 3.91)
aFor Canada, lag dsys with an "a" next to them represent the risk estimstes stsndsrdized to an approximate IQR of 5.1 ppb for a 1 -h max
increase in 03 concentrations.
bUnless noted, risk estimates standardized to 40 ppb for a 1 -h max increase in 03 concentrations.
1 For the U.S. in year round analyses, the investigators reported a 1.4% (95% CI: -0.9,
2 3.9%) increase in the PS model and 2.4% (95% CI: 0.0, 4.9%) increase in the NS model
3 in respiratory hospital admissions at lag 0-1 for a 40 ppb increase in 1-h max O3
4 concentrations with similar results for both models at lag 1 (Katsouyanni et al., 2009).
5 The distributed lag model provided results similar to those observed in the European
6 dataset with the PS model (1.1% [95% CI: -3.0, 5.3%]), but larger effects in the NS
7 model (3.3% [95% CI: 0.02, 6.8%]), which is consistent with the Canadian results. When
8 adjusting for PM10 using the U.S. data (i.e., every-6th-day PM data), results were
9 attenuated at lag 0-1 (PS: 0.6% [95% CI: -2.0, 3.3%]; NS: 1.4% [95% CI: -1.3, 4.2%])
10 which is consistent with the results presented for the European dataset. However, at lag 1,
11 U.S. risk estimates remained robust to the inclusion of PMi0 in copollutant models as was
12 observed in the Canadian and European datasets. Compared to the all-year analyses, the
13 investigators did not observe stronger associations in the summer-only analysis at either
14 lag 0-1 (-2.2%) or lag 1 (-2.8%) in both the PS and NS models (the study does not
15 present the extent of temporal smoothing used for these models).
16 Several additional multicity studies examined respiratory disease hospital admissions in
17 Canada and Europe. Cakmak et al. (2006b) evaluated the association between ambient O3
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1 concentrations and respiratory hospital admissions for all ages in 10 Canadian cities from
2 April 1993 to March 2000. The primary objective of this study was to examine the
3 potential modification of the effect of ambient air pollution on daily respiratory hospital
4 admissions by education and income using a time-series analysis conducted at the city-
5 level. The authors calculated a pooled estimate across cities for each pollutant using a
6 random effects model by first selecting the lag day with the strongest association from the
7 city-specific models. For O3, the mean lag day across cities that provided the strongest
8 association and for which the pooled effect estimate was calculated was 1.2 days. In this
9 study, all-year O3 concentrations were used in the analysis, and additional seasonal
10 analyses were not conducted. Cakmak et al. (2006b) reported a 4.4% increase (95% CI:
11 2.2, 6.5%) in respiratory hospital admissions for a 20 ppb increase in 24-h average O3
12 concentrations. The investigators only examined the potential effect of confounding by
13 other pollutants through the use of a multipollutant model (i.e., two or more additional
14 pollutants included in the model), which is difficult to interpret due to the potential
15 multicollinearity between pollutants. Cakmak et al. (2006b) also conducted an extensive
16 analysis of potential modifiers, specifically sex, educational attainment, and family
17 income, on the association between air pollution and respiratory hospital admissions.
18 When stratifying by sex, the increase in respiratory hospital admissions due to short-term
19 O3 exposure were similar in males (5.2% [95% CI: 3.0, 7.3%]) and females (4.2% [95%
20 CI: 1.8, 6.6%]). In addition, the examination of effect modification by income found no
21 consistent trend across the quartiles of family income. However, there was evidence that
22 individuals with an education level less than the 9th grade were disproportionately
23 affected by O3 exposure (4.6% [95% CI: 1.8, 7.5%]) compared to individuals that
24 completed grades 9-13 (1.7% [95% CI: -1.9, 5.3%]), some university or trade school
25 (1.4% [95% CI: -2.0, 5.1%]), or have a university diploma (0.66% [95% CI: -3.3, 4.7%]).
26 The association between O3 and individuals with an education level less than the 9th
27 grade was the strongest association across all of the pollutants examined.
28 A multicity study conducted in Europe by Biggeri et al. (2005) examined the association
29 between short-term O3 exposure and respiratory hospital admissions for all ages in four
30 Italian cities from 1990 to 1999. In this study, O3 was only measured during the warm
31 season (May-September). The authors examined associations between daily respiratory
32 hospital admissions and short-term O3 exposure at the city-level using a time-series
33 analysis. Pooled estimates were calculated by combining city-specific estimates using
34 fixed and random effects models. The investigators found no evidence of an association
35 between O3 exposure and respiratory hospital admissions in the warm season in both the
36 random (0.1% [95% CI: -5.2, 5.7%]; distributed lag 0-3) and fixed effects (0.1% [95%
37 CI: -5.2, 5.7%]; distributed lag 0-3) models for a 30 ppb increase in 8-h max O3
38 concentrations.
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1 Additional studies examined associations between short-term O3 exposure and respiratory
2 hospital admissions specifically in children. In a multicity study conducted in Canada,
3 Dales et al. (2006) examined the association between all-year ambient O3 concentrations
4 and neonatal (ages 0-27 days) respiratory hospital admissions in 11 Canadian cities from
5 1986 to 2000. The investigators used a statistical analysis approach similar to Cakmak et
6 al. (2006b) (i.e., time-series analysis to examine city-specific associations, and then a
7 random effects model to pool estimates across cities). The authors reported that for O3
8 the mean lag day across cities that provided the strongest association was 2 days. The
9 authors reported a 5.4% (95% CI: 2.9, 8.0%) increase in neonatal respiratory hospital
10 admissions for a 20 ppb increase in 24-h avg O3 concentrations at lag-2 days. The results
11 from Dales et al. (2006) provide support for the associations observed in a smaller scale
12 study that examined O3 exposure and pediatric respiratory hospital admissions in
13 New York state (Lin et al.. 2008a). Lin et al. (2008a) observed a positive association
14 between O3 and pediatric (i.e., <18 years) respiratory admissions at lag 2 (results not
15 presented quantitatively) in a two-stage Bayesian hierarchical model analysis of 11
16 geographic regions of New York from 1991 to 2001.
17 Overall, the evidence from epidemiologic studies continues to support an association
18 between short-term O3 exposure and respiratory-related hospital admissions, but it
19 remains unclear whether certain factors (individual- or population-level) modify this
20 association. Wong et al. (2009) examined the potential modification of the relationship
21 between ambient O3 (along with NO2, SO2, and PM10) and respiratory hospital
22 admissions by influenza intensity in Hong Kong for the period 1996 - 2002. Influenza
23 intensity was defined as a continuous variable using the proportion of weekly specimens
24 positive for influenza A or B instead of defining influenza epidemics. This approach was
25 used to avoid any potential bias associated with the unpredictable seasonality of influenza
26 in Hong Kong (Wong et al.. 2009). In models that examined the baseline effect (i.e.,
27 without taking into consideration influenza intensity) of short-term O3 exposure, the
28 authors found a 3.6% (95% CI: 1.9, 5.3%) and 3.2% (95% CI: 1.0, 5.4%) increase in
29 respiratory hospital admissions at lag 0-1 for a 30 ppb increase in 8-h max O3
30 concentrations for the all age and > 65 age groups, respectively. When examining
31 influenza intensity, Wong et al. (2009) reported that the association between short-term
32 exposure to O3 and respiratory hospital admissions was stronger with higher levels of
33 influenza intensity: additional increase in respiratory hospital admissions above baseline
34 of 1.4% (95% CI: 0.24, 2.6%) for all age groups and 2.4% (95% CI: 0.94, 3.8%) for those
35 65 and older when influenza activity increased from 0% to 10%. No difference in effects
36 was observed when stratifying by sex.
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Cause-Specific Respiratory Outcomes
1 In the 2006 O3 AQCD a limited number of studies were identified that examined the
2 effect of short-term O3 exposure on cause-specific respiratory hospital admissions. The
3 limited evidence "reported positive O3 associations with... asthma and COPD,
4 especially... during the summer or warm season" (U.S. EPA. 2006b). Of the studies
5 evaluated since the completion of the 2006 O3 AQCD, more have focused on identifying
6 whether O3 exposure is associated with specific respiratory-related hospital admissions,
7 including COPD, pneumonia, and asthma, but the overall body of evidence remains
8 small.
Chronic Obstructive Pulmonary Disease
9 Medina-Ramon et al. (2006) examined the association between short-term exposure to
10 ambient O3 and PM10 concentrations and Medicare hospital admissions among
11 individuals > 65 years of age for COPD in 35 cities in the U.S. for the years 1986-1999.
12 The cities included in this analysis were selected because they monitored PM10 on a daily
13 basis. In this study, city-specific results were obtained using a monthly time-stratified
14 case-crossover analysis. A meta-analysis was then conducted using random effects
15 models to combine the city-specific results. All cities measured O3 from May through
16 September, while only 16 of the cities had year-round measurements. The authors
17 reported a 1.6% increase (95% CI: 0.48, 2.9%) in COPD admissions for lag 0-1 in the
18 warm season for a 30 ppb increase in 8-h max O3 concentrations. When examining
19 single-day lags, stronger associations were observed for lag 1 (2.9% [95% CI: 1.8, 4.0%])
20 compared to lag 0 (-1.5% [95% CI: -2.7, -0.24%]). The authors found no evidence of
21 associations in cool season (-1.9% [95% CI: -3.6, -0.06%]; lag 0-1) or year round (0.24%
22 [95% CI: -0.78, 1.2%]; lag 0-1) analyses. In a copollutant model using warm season data,
23 the association between O3 and COPD hospital admissions was robust to the inclusion of
24 PM10 in the model (results not presented quantitatively). The authors conducted
25 additional analyses to examine potential modification of the warm season estimates for
26 O3 and COPD admissions by several city-level characteristics: percentage living in
27 poverty, emphysema mortality rate (as an indication of smoking), daily summer apparent
28 temperature, and percentage of households using central air conditioning. Of the city-
29 level characteristics examined, stronger associations were only reported for cities with a
3 0 larger variability in daily apparent summer temperature.
31 In a single-city study conducted in Vancouver from 1994-1998, a location with low
32 ambient O3 concentrations (Table 6-25), Yang et al. (2005b) examined the association
33 between O3 and COPD. Ozone was moderately inversely correlated with CO (r=-0.56),
34 NO2 (r=-0.32), and SO2 (r=-0.34), and weakly inversely correlated with PM10 (r=-0.09),
3 5 suggesting that the observed O3 effect is likely not only due to a positive correlation with
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1 other pollutants. Yang et al. (2005b) examined 1- to 7-day (e.g., (0-6 days) lagged
2 moving averages and observed an 8.8% (95% CI: -12.5, 32.6%) increase in COPD
3 admissions for lag 0-3 per 20 ppb increase in 24-h avg O3 concentrations. In two-
4 pollutant models at lag 0-3, O3 effect estimates were robust to the inclusion of NO2, SO2,
5 and PM10 in the model, but were increased slightly when adding CO (Figure 6-20; Table
6 6-28).
Pneumonia
7 In addition to COPD, Medina-Ramon et al. (2006) examined the association between
8 short-term exposure to ambient O3 and PMi0 concentrations and Medicare hospital
9 admissions among individuals > 65 years of age for pneumonia (ICD-9: 480-487). The
10 authors reported an increase in pneumonia hospital admissions in the warm season (2.5%
11 [95% CI: 1.6, 3.5%] for a 30 ppb increase in 8-h max O3 concentrations; lag 0-1). Similar
12 to the results observed for COPD hospital admissions, pneumonia hospital admissions
13 associations were stronger at lag 1 (2.6% [95% CI: 1.8, 3.4%]) compared to lag 0 (0.06%
14 [95% CI: -0.72, 0.78%]), and no evidence of an association was observed in the cool
15 season or year round. In two-pollutant models, the association between O3 exposure and
16 pneumonia hospital admissions was robust to the inclusion of PMi0 (results not presented
17 quantitatively). The authors also examined potential effect modification of the warm
18 season estimates for O3-related pneumonia hospital admissions, as was done for COPD,
19 by several city-level characteristics. Stronger associations were reported in cities with a
20 lower percentage of central air conditioning use. Across the cities examined, the
21 percentage of households having central air conditioning ranged from 6 to 93%. The
22 authors found no evidence of effect modification of the O3-pneumonia hospital admission
23 relationship when examining the other city-level characteristics.
24 Results from a single-city study conducted in Boston did not support the results presented
25 by Medina-Ramon et al. (2006). Zanobetti and Schwartz (2006) examined the association
26 of O3 and pneumonia Medicare hospital admissions for the period 1995-1999. Ozone was
27 weakly positively correlated with PM2 5 (r=0.20) and weakly inversely correlated with
28 black carbon, NO2, and CO (-0.25, -0.14, and -0.30, respectively). In an all-year analysis,
29 the investigators reported a 3.8% (95% CI: -7.9, -0.1%) decrease in pneumonia
30 admissions for a 20 ppb increase in 24-h average O3 concentrations at lag 0 and a 6.0%
31 (95% CI: -11.1, -1.4%) decrease for the average of lags 0 and 1. It should be noted that
32 the mean daily counts of pneumonia admissions was low for this study, ~14 admissions
33 per day compared to -271 admissions per day for Medina-Ramon et al. (2006). However,
34 in analyses with other pollutants Zanobetti and Schwartz (2006) did observe positive
3 5 associations with pneumonia hospital admissions, indicating that the low number of daily
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1 hospital admission counts probably did not influence the O3-pneumonia hospital
2 admissions association in this study. .
Asthma
3 There are relatively fewer studies that examined the association between short-term
4 exposure to O3 and asthma hospital admissions, presumably due to the limited power
5 given the relative rarity of asthma hospital admissions compared to ED or physician
6 visits. A study from New York City examined the association of 8-h max O3
7 concentrations with severe acute asthma admissions (i.e., those admitted to the Intensive
8 Care Unit [ICU]) during the warm season in the years 1999 through 2006 (Silverman and
9 Ito, 2010). In this study, O3 was moderately correlated with PM10 (r=0.59). When
10 stratifying by age, the investigators reported positive associations with ICU asthma
11 admissions for the 6-to 18-year age group (26.8% [95% CI: 1.4, 58.2%] for a 30 ppb
12 increase in maximum 8-h avg O3 concentrations at lag 0-1), but little evidence of
13 associations for the other age groups examined (<6 years, 19-49, 50+, and all ages).
14 However, positive associations were observed for each age-stratified group and all ages
15 for non-ICU asthma admissions, but again the strongest association was reported for the
16 6- to 18-years age group (28.2% [95% CI: 15.3, 41.5%]; lag 0-1). In two-pollutant
17 models, O3 effect estimates for both non-ICU and ICU hospital admissions remained
18 robust to adjustment for PM2 5. In an additional analysis, using a smooth function, the
19 authors examined whether the shape of the C-R curve for O3 and asthma hospital
20 admissions (i.e., both general and ICU for all ages) is linear. To account for the potential
21 confounding effects of PM25, Silverman and Ito (2010) also included a smooth function
22 of PM2 5 lag 0-1. When comparing the curve to a linear fit line the authors found that the
23 linear fit is a reasonable approximation of the concentration-response relationship
24 between O3 and asthma hospital admissions around and below the level of the current
25 NAAQS (Figure 6-15).
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Ozone: All Ages
cc
cc
o>
o
20
I
40
l
60
Ozone
80
100
Source: Used with permission from American Academy of Allergy, Asthma & Immunology (Silverman and Ito. 2010).
The average of 0 day and 1 day lagged 8-h ozone was used in a two-pollutant model with PM2 5 lag 0-1, adjusting for temporal
trends, day of the week, and immediate and delayed weather effects. The solid lines are smoothed fit data, with long broken lines
indicating 95% confidence bands. The density of lines at the bottom of the figure indicates sample size.
Figure 6-15 Estimated relative risks (RRs) of ozone-related asthma hospital
admissions allowing for possible nonlinear relationships using
natural splines.Averting Behavior
1 The studies discussed above have found consistent positive associations between short-
2 term O3 exposure and respiratory-related hospital admissions, however, the strength of
3 these associations may be underestimated due to the studies not accounting for averting
4 behavior. As discussed in Section 4.6.4, recent studies by Neidell (2009) and Neidell and
5 Kinney (2010) conducted in Souther California demonstrate that controlling for
6 avoidance behavior increases O3 effect estimates for respiratory hospital admissions,
7 specifically for children and older adults. These studies show that on days where no
8 public alert was issued warning of high O3 concentrations there was an increase in asthma
9 hospital admissions. Although only a few epidemiologic studies have examined averting
10 behavior and these studies are limited to asthma hospital admissions, they do provide
11 preliminary evidence indicating that epidemiologic studies may underestimate
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1 associations between O3 exposure and health effects by not accounting for behaviorial
2 modification when public health alerts are issued.
6.2.7.3 Emergency Department Visit Studies
3 Overall, relatively fewer studies have examined the association between short-term O3
4 exposure and respiratory-related ED visits, compared to hospital admissions. In the 2006
5 O3 AQCD, positive, but inconsistent, associations were observed between O3 and
6 respiratory-related ED visits with effects generally occurring during the warm season.
7 Since the completion of the previous AQCD, larger studies have been conducted, in
8 terms of sample size, study duration, and in some cases multiple cities, to examine the
9 association between O3 and ED visits for all respiratory diseases, COPD, and asthma.
Respiratory Disease
10 A large single-city study conducted in Atlanta, by Tolbert et al. (2007). and subsequently
11 reanalyzed by Darrow et al. (20 lib), provides evidence for an association between short-
12 term exposures to ambient O3 concentrations and respiratory ED visits. Tolbert et al.
13 (2007) examined the association between air pollution, both gaseous pollutants and PM
14 and its components, and respiratory disease ED visits in all ages from 1993 to 2004. The
15 correlations between O3 and the other pollutants examined ranged from 0.2 for CO and
16 SO2 to 0.5-0.6 for the PM measures. Using an a priori average of lags 0-2 for each air
17 pollutant examined, the authors reported a 3.9% (95% CI: 2.7, 5.2%) increase in
18 respiratory ED visits for a 30 ppb increase in 8-h max O3 concentrations during the warm
19 season [defined as March-October in Darrow et al. (20lib)]. In copollutant models, the
20 O3 associations with respiratory ED visits remained robust with CO, NO2, and PM10
21 (results not presented quantitatively).
22 Darrow et al. (20 lib) examined the same data as Tolbert et al. (2007). but explored
23 whether differences exist in the association between O3 exposure and respiratory-related
24 ED visits depending on the exposure metric used (i.e., 8-h max, 1-h max, 24-h average,
25 commuting period [7:00 a.m. to 10:00 a.m.; 4:00 p.m. to 7:00 p.m.], day-time [8:00 a.m.
26 to 7:00 p.m.] and night-time [12:00 a.m. to 6:00 a.m.]). To examine the association
27 between the various O3 exposure metrics and respiratory ED visits, the authors used a
28 time-stratified case-crossover approach, selecting control days as those days within the
29 same calendar month and maximum temperature as the case day. Darrow et al. (20 lib)
30 found at lag 1, the results were somewhat variable across exposure metrics. The strongest
31 associations with respiratory ED visits were found when using the 8-h max, 1-h max, and
32 day-time exposure metrics with weaker associations using the 24-h avg and commuting
Draft - Do Not Cite or Quote 6-124 September 2011
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1
2
period exposure metrics; a negative association was observed when using the night-time
exposure metric (Figure 6-16).
1.03 T
0
y 1.02
Ss
& 8- 1.01
je o
* -C
9
a. 0^g
Partial
Spearman r.
1 0.95 0.93 0.83 078 0.04
• XBiujq-g
>,
a
E
v-
commute -
4)
^-
CM
<->
_g>
"E
Source: Used with permission from Nature Publishing Group (Darrow et al.. 2011 b).
Figure 6-16 Risk ratio for respiratory ED visits and different ozone exposure
metrics in Atlanta from 1993-2004.
4
5
6
1
8
9
10
11
12
13
In an additional study conducted in 6 Italian cities, Orazzo et al. (2009) examined
respiratory ED visits for ages 0-2 years in 6 Italian cities from 1996 to 2000. However,
instead of identifying respiratory ED visits using the traditional approach of selecting
ICD codes as was done by Tolbert et al. (2007) and Darrow et al. (20lib). Orazzo et al.
(2009) used data on wheeze extracted from medical records as an indicator of lower
respiratory disease. This study examined daily counts of wheeze in relation to air
pollution using a time-stratified case-crossover approach in which control days were
matched on day of week in the same month and year as the case day. The authors found
no evidence of an association between 8-h max O3 concentrations and respiratory ED
visits in children aged 0-2 years in models that examined both single-day lags and
moving averages of lags from 0-6 days in year-round and seasonal analyses (i.e., warm
Draft - Do Not Cite or Quote
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1 and cool seasons). In all-year analyses, the percent increase in total wheeze ranged from -
2 1.4% to -3.3% for a 0-1 to 0-6 day lag, respectively.
COPD
3 Stieb et al. (2009) also examined the association between short-term O3 exposure and
4 COPD ED visits in 7 Canadian cities. Across cities, in an all-year analysis, O3 was found
5 to be positively associated with COPD ED visits (4.0% [95% CI: -0.54, 8.6%] at lag 2 for
6 a 20 ppb increase in 24-h avg O3 concentrations). In seasonal analyses, larger effects
7 were observed between O3 and COPD ED visits during the warm season (i.e., April-
8 September) 6.8% [95% CI: 0.11, 13.9%] (lag day not specified); with no associations
9 observed in the winter season. Stieb et al. (2009) also examined associations between
10 respiratory-related ED visits, including COPD, and air pollution at sub-daily time scales
11 (i.e., 3-h avg of ED visits versus 3-h avg pollutant concentrations) and found no evidence
12 of consistent associations between any pollutant and any respiratory outcome.
13 In a single-city study, Arbex et al. (2009) examined the association between COPD and
14 several ambient air pollutants, including O3, in Sao Paulo, Brazil for the years 2001-2003
15 for individuals over the age of 40. Associations between O3 exposure and COPD ED
16 visits were examined in both single-day lag (0-6 days) and polynomial distributed lag
17 models (0-6 days). In all-year analyses, O3 was not found to be associated with an
18 increase in COPD ED visits (results not presented quantitatively). The authors also
19 conducted stratified analyses to examine the potential modification of the air pollutant-
20 COPD ED visits relationship by age (e.g., 40-64, >64) and sex. In these analyses O3 was
21 found to have an increase in COPD ED visits for women, but not for men or either of the
22 age groups examined.
Asthma
23 In a study of 7 Canadian cities, Stieb et al. (2009) also examined the association between
24 exposure to air pollution (i.e., CO, NO2, O3, SO2, PM10, PM2.5, and O3) and asthma ED
25 visits. Associations between short-term O3 exposure and asthma ED visits were examined
26 at the city level and then pooled using either fixed or random effects models depending
27 on whether heterogeneity among effect estimates was found to be statistically significant.
28 Across cities, in an all-year analysis, the authors found that short-term O3 exposure was
29 associated with a positive increase (3.5% [95% CI: 0.33, 6.8%] at lag 2 for a 20 ppb
30 increase in 24-h avg O3 concentrations) in asthma ED visits. The authors did not present
31 the results from seasonal analyses for asthma, but state that no associations were
32 observed between any pollutant and respiratory ED visits in the winter season. As stated
33 previously, in analyses of 3-h avg O3 concentrations, the authors observed no evidence of
34 consistent associations between any pollutant and any respiratory outcome, including
Draft - Do Not Cite or Quote 6-126 September 2011
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1 asthma. A single-city study conducted in Alberta, Canada (Villeneuve et al., 2007) from
2 1992-2002 among individuals two years of age and older provides additional support for
3 the findings from Stieb et al. (2009). but also attempts to identify those lifestages (i.e., 2-
4 4, 5-14, 15-44, 45-64, 65-74, or 75+) most susceptible to O3-induced asthma ED visits. In
5 a time-referent case-crossover analysis, Villeneuve et al. found an increase in asthma ED
6 visits in an all-year analysis across all ages (12.0% [95% CI: 6.8, 17.2] for a 30 ppb
7 increase in max 8-h avg O3 concentrations at lag 0-2) with associations being stronger
8 during the warmer months (19.0% [95% CI: 11.9, 28.1]). When stratifying by age, the
9 strongest associations were observed in the warm season for individuals 5-14 (28.1%
10 [95% CI: 11.9, 45.1]; lag 0-2) and 15-44 (19.0% [95% CI: 8.5, 31.8]; lag 0-2). These
11 associations were not found to be confounded by the inclusion of aeroallergens in age-
12 specific models.
13 Several additional single-city studies have also provided evidence of an association
14 between asthma ED visits and ambient O3 concentrations. Ito et al. (2007b) examined the
15 association between short-term exposure to air pollution and asthma ED visits for all ages
16 in New York City from 1999 to 2002. Ito et al. (2007b) used three different weather
17 models with varying extent of smoothing to account for temporal relationships and
18 multicollinearity among pollutants and meteorological variables (i.e., temperature and
19 dew point) to examine the effect of model selection on the air pollutant-asthma ED visit
20 relationship. When examining O3, the authors reported a positive association with asthma
21 ED visits, during the warm season across the models (ranging from 8.6 to 16.9%) and an
22 inverse association in the cool season (ranging from -23.4 to -25.1%), at lag 0-1 for a 30
23 ppb increase in 8-h max O3 concentrations. Using a simplified version of the weather
24 model used in NMMAPS analyses (i.e., terms for same-day temperature and 1-3 day
25 average temperature), Ito et al. (2007b) found that O3 effects were not substantially
26 changed in copollutant models with PM2 5, NO2, SO2, and CO during the warm season
27 (Figure 6-19; Table 6-27).
Draft - Do Not Cite or Quote 6-127 September 2011
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Ozone Warm Season
40 50 60 70
Concentration (ppb)
80
Source: Used with permission from American Thoracic Society (Strickland et al.. 2010).
The reference for the rate ratio is the estimated rate at the 5th percentile of the pollutant concentration. Estimates are presented
for the 5th percentile through the 95th percentile of pollutant concentrations due to instability in the dose-response estimates at the
distribution tails.
Figure 6-17 Loess dose-response estimates and twice-standard error estimates
from generalized additive models for associations between 3-day
avg ozone concentrations and ED visits for pediatric asthma.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
Strickland et al. (2010) examined the association between O3 exposure and pediatric
asthma ED visits (ages 5-17 years) in Atlanta between 1993 and 2004 using the same air
quality data as Darrow et al. (20lib) and Tolbert et al. (2007). However, unlike Darrow
et al. (20lib) and Tolbert et al. (2007). which used single centrally located monitors or an
average of monitors, respectively, Strickland et al. (2010) used population-weighting to
combine daily pollutant concentrations across monitors. In this study, the authors
developed a statistical model using hospital-specific time-series data that is essentially
equivalent to a time-stratified case-crossover analysis (i.e., using interaction terms
between year, month, and day-of-week to mimic the approach of selecting referent days
within the same month and year as the case day). The authors observed a 6.4% (95% CI:
3.2, 9.6%) increase in ED visits for a 30 ppb increase in 8-h max O3 concentrations at lag
0-2 in an all-year analysis. In seasonal analyses, stronger associations were observed
during the warm season (i.e., May-October) (8.4% [95% CI: 4.4, 12.7%]; lag 0-2) than
the cold season (4.5% [95% CI: -0.82, 10.0%]; lag 0-2). Strickland et al. (2011)
confirmed these findings in an additional analysis using the same dataset, and found that
the metric used to assign exposure (i.e., centrally located monitor, unweighted average
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1 across monitors, and population-weighted average across monitors) did not influence
2 pediatric asthma ED visit risk estimates for spatially homogeneous pollutants such as O3.
3 In copollutant analyses, Strickland et al. (2010) found that O3 effect estimates were not
4 substantially changed when controlling for other pollutants (CO, NO2, PM2.5 elemental
5 carbon, PM2 5 sulfate) (results not presented quantitatively). The authors also examined
6 the C-R relationship between O3 exposure and pediatric asthma ED visits and found that
7 both quintile and loess dose-response analyses (Figure 6-17) suggest that there are
8 elevated associations with O3 at concentrations as low as 30 ppb. These dose-response
9 analyses do not provide evidence of a threshold level.
10 In a single-city study conducted on the West coast, Mar and Koenig (2009) examined the
11 association between O3 exposure and asthma ED visits (ICD-9 codes_ 493-493.9) for
12 children (< 18) and adults (> 18) in Seattle, WA from 1998 to 2002. Of the total number
13 of visits over the study duration, 64% of visits in the age group < 18 comprised boys, and
14 70% of visits in the > 18 age group comprised females. Mar and Koenig (2009)
15 conducted a time-series analysis using both 1-h max and max 8-h avg O3 concentrations.
16 Although a similar pattern of associations was observed using both metrics, only those
17 results using the max 8-h avg O3 metric are discussed here since they are more applicable
18 to the current O3 NAAQS. Mar and Koenig (2009) presented results for single day lags of
19 0 to 5 days, but found consistent positive associations across individual lag days which
20 supports the findings from the studies discussed above that examined multi-day
21 exposures. For children, consistent positive associations were observed across all lags,
22 ranging from a 19.1-36.8% increase in asthma ED visits for a 30 ppb increase in 8-h max
23 O3 concentrations with the strongest associations observed at lag 0 (33.1% [95% CI: 3.0,
24 68.5]) and lag 3 (36.8% [95% CI: 6.1, 77.2]) (Figure x). O3 was also found to be
25 positively associated with asthma ED visits for adults at all lags, ranging from 9.3-26.0%,
26 except at lag 0 (Figure 6-18). The slightly different lag times for children and adults
27 suggest that children may be more immediately responsive to O3 exposures than adults
28 (Mar and Koenig. 2009).
Draft - Do Not Cite or Quote 6-129 September 2011
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LagO
Lagl
Lag 2
Lag3
Lag 4
LagS
Black = Children
Red = Adults
0.80 1.00
1.20 1.40
Relative Risk
1.60
1.80
Adapted from Mar and Koenig (2009).
Figure 6-18 Relative risk of asthma ED visits children and adults for a 30 ppb
increase in max 8-h avg O3 concentrations in Seattle, WA, 1998-
2002.
1
2
3
4
5
6
7
Respiratory Infection
Although an increasing number of studies have examined the association between O3
exposure and cause-specific respiratory ED visits this trend has not included an extensive
examination of the association between O3 exposure and respiratory infection ED visits.
Stieb et al. (2009) also examined the association between short-term O3 exposure and
respiratory infection ED visits in 7 Canadian cities. In an all-year analysis, there was no
evidence of an association between O3 exposure and respiratory infection ED visits at all
lags examined (i.e., 0, 1, and 2). Across cities, respiratory infections comprised the single
largest diagnostic category, approximately 32%, of all the ED visits examined, which
also included myocardial infarction, heart failure, dysrhythmia, asthma, and COPD.
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6.2.7.4 Outpatient and Physician Visit Studies
1 Several studies have examined the association between ambient O3 concentrations and
2 physician or outpatient (non-hospital, non-ED) visits for acute conditions in various
3 geographic locations. Burra et al. (2009) examined asthma physician visits among
4 patients aged 1-17 and 18-64 years in Toronto, Canada from 1992 to 2001. The authors
5 found little or no evidence of an association between asthma physician visits and O3;
6 however, seasonal analyses were not conducted. It should be noted that in this study,
7 most of the relative risks for O3 were less than one and statistically significant, perhaps
8 indicating an inverse correlation with another pollutant or an artifact of the strong
9 seasonality of asthma visits. Villeneuve et al. (2006b) also focused on physician visits to
10 examine the effect of short-term O3 exposure on allergic rhinitis among individuals aged
11 65 or older in Toronto from 1995 to 2000. The authors did not observe any evidence of
12 an association between allergic rhinitis physician visits and ambient O3 concentrations in
13 single-day lag models in an all-year analysis (results not presented quantitatively).
14 In a study conducted in Atlanta, Sinclair et al. (2010) examined the association of acute
15 asthma and respiratory infection (e.g., upper respiratory infections and lower respiratory
16 infections) outpatient visits from a managed care organization with ambient O3
17 concentrations as well as multiple PM size fractions and species from August 1998
18 through December 2002. The authors separated the analysis into two time periods (the
19 first 25 months of the study period and the second 28 months of the study period), in
20 order to compare the air pollutant concentrations and relationships between air pollutants
21 and acute respiratory visits for the 25-month time-period examined in Sinclair et al.
22 (2004) to an additional 28-month time-period of available data from the Atlanta Aerosol
23 Research Inhalation Epidemiology Study (ARIES). The authors found little evidence of
24 an association between O3 and asthma visits, for both children and adults, or respiratory
25 infection visits in all-year analyses and seasonal analyses. For example, a slightly
26 elevated relative risk (RR) for childhood asthma visits was observed during the 25-month
27 period in the cold season (RR: 1.12 [95% CI: 0.86, 1.41]; lag 0-2 for a 30 ppb increase in
28 8-h max O3), but not in the warm season (RR: 0.97 [95% CI: 0.86, 1.10]; lag 0-2). During
29 the 28-month period at lag 0-2, a slightly larger positive effect was observed during the
30 warm season (RR: 1.06 [95% CI: 0.97, 1.17]), compared to the cold season (RR: 1.03
31 [95% CI: 0.87, 1.21]). Overall, these results contradict those from Strickland et al. (2010)
32 discussed above. Although the mean number of asthma visits and O3 concentrations in
33 Sinclair et al. (2010) and Strickland et al. (2010) are similar the difference in results
34 between the two studies could be attributed to the severity of O3-induced asthma
35 exacerbations (i.e., more severe symptoms requiring a visit to a hospital) and behavior,
36 such as delaying a visit to the doctor for less severe symptoms.
Draft - Do Not Cite or Quote 6-131 September 2011
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6.2.7.5 Summary
1 The results of the recent studies evaluated largely support the conclusion of the 2006 O3
2 AQCD. While fewer studies were published overall since the previous review, several
3 multicity studies (e.g., Cakmak et al.. 2006b: Dales et al.. 2006) and a multi-continent
4 study (Katsouvanni et al., 2009) provide supporting evidence for an association between
5 short-term O3 exposure and an increase in respiratory-related hospital admissions and ED
6 visits. Collectively, in the studies evaluated, both single-city and multicity, there is
7 continued evidence for increases in both hospital admissions and ED visits when
8 examining all respiratory outcomes combined. Additionally, new studies support an
9 association between short-term O3 exposure and asthma (Strickland et al.. 2010; Stieb et
10 al.. 2009) and COPD (Stieb et al.. 2009; Medina-Ramon et al.. 2006) hospital admissions
11 and ED visits, with more limited evidence for pneumonia hospital admissions and ED
12 visits (Medina-Ramon et al., 2006; Zanobetti and Schwartz. 2006). In seasonal analyses,
13 stronger associations were observed in the warm season or summer months compared to
14 the cold season, particularly for asthma (Strickland et al.. 2010; Ito et al.. 2007b) and
15 COPD (Medina-Ramon et al.. 2006) (Figure 6-19; Table 6-27), which is consistent with
16 the conclusions of the 2006 O3 AQCD. There is also continued evidence that children are
17 particularly susceptible to O3-induced respiratory effects (Silverman and Ito. 2010;
18 Strickland et al.. 2010; Mar and Koenig. 2009; Villeneuve et al.. 2007; Dales et al..
19 2006). Although the collective evidence indicates a consistent positive association
20 between O3 exposure and respiratory-related hospital admissions and ED visits, the
21 magnitude of these associations may be underestimated due to behavioral modification in
22 response to forecasted air quality (Neidell and Kinney. 2010; Neidell. 2009)
23 (Section 4.6.4).
24 Additional studies that focused on respiratory-related outpatient or physician visits found
25 no evidence of an association with short-term O3 exposure, but this could be attributed to
26 the severity of O3-induced respiratory effects requiring more immediate treatment or
27 behavioral factors that result in delayed visits to a physician.
28 The studies that examined the potential confounding effects of copollutants found that O3
29 effect estimates remained relatively robust upon the inclusion of PM and gaseous
30 pollutants in two-pollutant models (Figure 6-19; Table 6-27), including (Strickland et al..
31 2010; Tolbert et al.. 2007; Medina-Ramon et al.. 2006). which did not present results
32 quantitatively. These findings are consistent with the studies evaluated in the 2006 O3
33 AQCD (U.S. EPA. 2006b) (Figure 7-12, p. 7-80) which found O3 respiratory hospital
34 admissions risk estimates remained robust to the inclusion of PM in copollutant models.
Draft - Do Not Cite or Quote 6-132 September 2011
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Study
Visit Type
Age
Wong etal. (2009; 196722]
Cakmak etal. (2006; 93272)
Dales etal. (2006; 90744]
Orazzoetal.(2009 202800)a
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Darrow etal. (2009; 202800]
Tol be rt etal. (2007,090316)
Biggeri etal. (2005 87395]c
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Katsouyanni et al. 2009; 199899
Stiebetal. (2009: 195858]
Villeneuve et al. (2007; 195859]
Strickland etal. (2010; 624878]
Silverman and Ito 2010; 386252]c
Ito etal. (2007; 156594]
Villeneuve et al. (2007; 195859)
Hong Kong
10 Canadian cities
11 Canadian cities
6 Italian cities
APHENA-Europe
APHENA-U.S.
APHENA-Canada
b APHENA-Canada
Atlanta
Atlanta
8 Italian cities
APHENA-Europe
APHENA-U.S.
APHENA-Canada
b APHENA-Canada
7 Canadian Cities
Alberta, CAN
Atlanta
New York
New York
Alberta. CAN
Mar and Koenig [2009; 594410] Seattle,' WA
Strickland et al. (2010; 624878] Atlanta
Silverman and Ito (2010; 386252)d New York
Mar and Koenig (2009; 594410]
Ito etal. (2007; 156594]
Villeneuve et al. (2007; 195859)
Strickland etal. (2010; 624878]
Stiebetal. (2009; 195858]
Medina-Ramon eta. (2006; 8772
Yang etal. (2006; 90184)
Stiebetal. (2009; 195858]e
Medina-Ramon et a. (2006; 8772
Medina-Ramon eta. (2006; 8772
Seattle, WA
New York
Alberta, CAN
Atlanta
7 Canadian Cities
] 36 U.S. cities
Vancouver
7 Canadian Cities
36 U.S. cities
36 U.S. cities
Zanobetti and Schwartz (2006; 90195] Boston
Medina-Ramon eta . 2006; 8772
Medina-Ramon eta. 2006; 8772
Medina-Ramon eta . 2006; 8772
36 U.S. cities
36 U.S. cities
36 U.S. cities
HA
HA
HA
ED
HA
HA
HA
HA
ED
ED
HA
HA
HA
HA
HA
ED
ED
ED
HA
ED
ED
ED
ED
HA
ED
ED
ED
ED
ED
HA
HA
ED
HA
HA
HA
HA
HA
HA
All
All
0-27 days
0-2
65+
65+
65+
65+
All
All
All
65+
65+
65+
65+
All
>2
Children
All
All
>2
18+
Children
6-18
<18
All
>2
Children
All
65+
65+
All
65+
65+
65+
65+
65+
65+
0-1
1.2
2
0-6
0-1
0-1
DLjO-2]
DLJO-2]
1
0-2
0-3
0-1
0-1
DLjO-2]
DLJO-2]
2
0-2
0-2
0-1
0-1
0-2
2
0-2
0-1
0
0-1
0-2
0-2
2
0-1
0-3
NR
0-1
0-1
0-1
0-1
0-1
0-1
Respiratory
Asthma
25 -20 -15 -10 -5 0 5 10 15 20 25 30 35 40
3 Wheeze used as indicator of lower respiratory disease.
b APHENA-Canada results standardized to approximate IQR of 5.1 ppb for 1 h max O3 concentrations.
0 Study included 8 cities; but of those 8, only 4 had O3 data.
d non-ICU effect estimates.
e The study did not specify the lag day of the summer season estimate.
Effect estimates are for a 20 ppb increase in 24 hours; 30 ppb increase in 8-h max; and 40 increase in 1-h max ozone
concentrations. HA=hospital admission; ED=emergency department. Black=AII-year analysis; Red=Summer only analysis;
Blue=Winter only analysis.
Figure 6-19 Percent increase in respiratory-related hospital admission and ED
visits in studies that presented all-year and/or seasonal results.
Draft - Do Not Cite or Quote
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September 2011
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Table 6-27 Corresponding Effect Estimates for Figure 6-19
Study
ED Visit or
Hospital
Admission
Location
Age
Lag
Avg Time
% Increase
(95% Cl)
Respiratory
All-year
Wona et al. (2009)
Cakmaketal. (2006b)
Dales et al. (2006)
Orazzoetal. (2011 b)a
Katsouyanni et al. (2009)
Hospital Admission
Hospital Admission
Hospital Admission
ED Visit
Hospital Admission
Hong Kong
10 Canadian cities
11 Canadian cities
6 Italian cities
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
All
All
0-27 days
0-2
65+
65+
65+
65+
0-1
1.2
2
0-6
0-1
0-1
DL(0-2)
DL(0-2)b
8-h max
24-h avg
24-h avg
8-h max
1-hmax
1-hmax
1-hmax
1-hmax
3.58(1.90,5.29)
4.38(2.19,6.46)
5.41 (2.88, 7.96)
-3.34 (-11. 2, 5.28)
1.58 (-1.71, 4.15)
2.38 (0.00, 4.89)
20.4 (4.07, 40.2)
2.4(0.51,4.40)
Warm
Darrowetal. (2011 b)
Tolbert et al. (2007)
Bidden et al. (2005)°
Katsouyanni et al. (2009)
ED Visit
ED Visit
Hospital Admission
Hospital Admission
Atlanta
Atlanta
8 Italian cities
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
All
All
All
65+
65+
65+
65+
1
0-2
0-3
0-1
0-1
DL(0-2)
DL(0-2)b
8-h max
8-h max
8-h max
1-h max
1-h max
1-h max
1-h max
2.08(1.25,2.91)
3.90 (2.70, 5.20)
0.06 (-5.24, 5.66)
0.24 (-3.32, 3.91)
2.14 (-0.63, 4.97)
37.1(11.5,67.5)
4.1(1.40,6.80)
Asthma
All-year
Stiebetal. (2009)
Villeneuve et al. (2007)
Strickland etal. (2010)
ED Visit
ED Visit
ED Visit
7 Canadian cities
Alberta, CAN
Atlanta
All
>2
Children
2
0-2
0-2
24-h avg
8-h max
8-h max
3.48 (0.33, 6.76)
11.9(6.8,17.2)
6.38(3.19,9.57)
Warm
Silverman and Ito (2010)"
Itoetal. (2007b)
Villeneuve et al. (2007)
Mar and Koenid (2009)
Strickland etal. (2010)
Silverman and Ito (2010)"
Mar and Koenid (2009)
Hospital Admission
ED Visit
ED Visit
ED Visit
ED Visit
Hospital Admission
ED Visit
New York
New York
Alberta, CAN
Seattle, WA
Atlanta
New York
Seattle, WA
All
All
>2
18+
Children
6-18
<18
0-1
0-1
0-2
2
0-2
0-1
0
8-h max
8-h max
8-h max
8-h max
8-h max
8-h max
8-h max
12.5(8.27,16.7)
16.9(10.9,23.4)
19.0(11.9,28.1)
19.1(3.00,40.5)
8.43(4.42,12.7)
28.2(15.3,41.5)
33.1(3.00,68.5)
Cold
Itoetal. (2007b)
Villeneuve et al. (2007)
Strickland etal. (2010)
ED Visit
ED Visit
ED Visit
New York
Alberta, CAN
Atlanta
All
>2
Children
0-1
0-2
0-2
8-h max
8-h max
8-h max
-23.4 (-27.3, -19.3)
8.50(0.00,17.2)
4.52 (-0.82, 10.1)
COPD
All-year
Stiebetal. (2009)
Medina-Ramon et al. (2006)
Yand et al. (2005b)
ED Visit
Hospital Admission
Hospital Admission
7 Canadian cities
36 U.S. cities
Vancouver
All
65+
65+
2
0-1
0-3
24-h avg
8-h max
24-h avg
4.03 (-0.54, 8.62)
0.24 (-0.78, 1.21)
8.80 (-12.5, 32.6)
Warm
Stiebetal. (2009)e
Medina-Ramon et al. (2006)
ED Visit
Hospital Admission
7 Canadian cities
36 U.S. cities
All
65+
NR
0-1
24-h avg
8-h max
6.76(0.11,13.9)
1.63(0.48,2.85)
Cold
Medina-Ramon et al. (2006)
Hospital Admission
36 U.S. cities
65+
0-1
8-h max
-1.85 (-3.60, -0.06)
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6-134
September 2011
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Study
ED Visit or
Hospital
Admission
Location
Age Lag Avg Time
% Increase
(95% Cl)
Pneumonia
All-year
Zanobetti and Schwartz (2006)
Medina-Ramon et al. (2006)
Hospital Admission
Hospital Admission
Boston
36 U.S. cities
65+ 0-1 24-h avg
65+ 0-1 8-h max
-5.96 (-11.1, -1.36)
1.81 (-0.72, 4.52)
Warm
Medina-Ramon et al. (2006)
Hospital Admission
36 U.S. cities
65+ 0-1 8-h max
2.49(1.57,3.47)
Cold
Medina-Ramon et al. (2006)
Hospital Admission
36 U.S. cities
65+ 0-1 8-h max
-4. 88 (-6. 59, -3.14)
"Wheeze used as indicator of lower respiratory disease.
bAPHENA-Canada results standardized to approximate IQRof 5.1 ppbfor 1-h max 03 concentrations.
°Study included 8 cities, but of those 8 only 4 had 03 data.
dNon-ICU effect estimates.
The study did not specify the lag day of the summer season estimate.
Study Location Age Lag Copollutant
Respiratory
Katsouyanni et al. (2009; 199899)a APHENA-U.S. 65+ 1
PM10
APHENA-Europe
PM10
c
pwiin
c PM10 —
COPD
fengetal. (2006;90184)a .an^uv.i C. I 0 „ <
Al 1-Year
«-
-•—
'
J°' *
Asthma
Itoetal. (2007;156594)b New York All 0-1
CO
NO2
SO2
PM2.5
Summer
-10 -50 5 10 15 20 25 30
% Increase
Effect estimates are for a 20 ppb increase in 24 hours; 30 ppb increase in 8-h max; and 40 ppb increase in 1-h max ozone
concentrations. An "a" represents studies that examined hospital admissions, "b" represents a study that examined ED visits, and "c"
represents risk estimates from APHENA-Canada standardized to an approximate IQR of 5.1 ppb for a 1-h max increase in ozone
concentrations. Black = results from single-pollutant models; Red = results from copollutant models with PM10or PM2.s; Yellow =
results from copollutant models with CO; Blue = results from copollutant models with NO2; Green = results from copollutant models
with SO2.
Figure 6-20 Percent increase in respiratory-related hospital admissions and ED
visits for studies that presented single and copollutant model
results.
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6-135
September 2011
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Table 6-28 Corresponding effect
Study3 Location Visit Type Age
estimates for Figure 6-20
Lag Copollutant
% Increase (95% Cl)
All-year
Respiratory
Katsouyanni et al. (2009) APHENA-U.S.
APHENA-Europe
APHENA-Canada
HA 65+ 1
PM,0
PM,0
PM,0
PM,0
2.62 (0.63, 4.64)
2.14 (-0.08, 4.40)
2.94(1.02,4.89)
2.38 (0.08, 4.64)
5.54 (-0.94, 12.4)
0.69 (-0.12, 1.50)b
5. 13 (-6.62, 18.6)
0.64 (-0.87, 2.20)b
COPD
Yang et al. (2005b) Vancouver
HA 65+ 0-3
CO
N02
S02
PM,0
8.80 (-12.5, 32.6)
22.8 (-2.14, 50.7)
11.1 (-10.4, 37.6)
13.4 (-8.40, 40.2)
11.1 (-8.40, 37.6)
Summer
Asthma
Itoetal. (2007b) New York
ED All 0-1
CO
N02
S02
PM2,
16.9(10.9,23.4)
18.1(12.1,24.5)
10.2(4.29,16.4)
13.1(7.16,19.5)
12.7(6.37,19.3)
"Averaging times: Katsouyanni et al. (2009) = 1-h max; Yang et al. (2005b) = 24-h avg; and Ito et al. (2007b) = 8-h max.
bRisk estimates standardized to an approximate IQRof 5.1 ppbfora 1-h max increase in 03 concentrations.
1 Additionally, a preliminary examination of the C-R relationship found no evidence of a
2 threshold between short-term O3 exposure and pediatric asthma ED visits (Silverman and
3 Ito. 2010; Strickland et al.. 2010). Overall, the new body of evidence supports an
4 association between short-term O3 exposure and respiratory-related hospital admissions
5 and ED visits, with additional evidence for stronger associations during the warm season
6 for specific respiratory outcomes such as asthma and COPD.
6.2.8 Respiratory Mortality
7 The epidemiologic, controlled human exposure, and toxicological studies discussed
8 within this section (Section 6.2) provides evidence for multiple respiratory effects in
9 response to short-term O3 exposure. Additionally, the evidence from experimental studies
10 indicates multiple potential pathways of O3-induced respiratory effects, which support the
11 continuum of respiratory effects that could potentially result in respiratory-related
12 mortality. The 2006 O3 AQCD found inconsistent evidence for an association between
13 short-term O3 exposure and respiratory mortality (U.S. EPA. 2006b). Although some
14 studies reported a strong positive association between O3 exposure and respiratory
Draft - Do Not Cite or Quote 6-136 September 2011
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1 mortality, additional studies reported a small association or no association. Recent
2 multicity studies found consistent positive associations between short-term O3 exposure
3 and respiratory mortality, specifically during the summer months.
4 The APHENA study, described earlier in Section 6.2.7.2, (Katsouyanni et al.. 2009) also
5 examined associations between short-term O3 exposure and mortality and found
6 consistent positive associations for respiratory mortality in all-year analyses with stronger
7 associations in analyses restricted to the summer season. Additional multicity studies
8 from the U.S. (Zanobetti and Schwartz. 2008b). Europe (Samoli et al.. 2009). Italy
9 (Stafoggia et al.. 2010). and Asia (Wong et al.. 2010) that conducted summer season
10 and/or all-year analyses provide additional support for an association between short-term
11 O3 exposure and respiratory mortality (Figure 6-37).
12 Of the studies evaluated, only the APHENA study (Katsouyanni et al.. 2009) and the
13 Italian multicity study (Stafoggia et al.. 2010) conducted an analysis of the potential for
14 copollutant confounding of the O3-respiratory mortality relationship. In the APHENA
15 study, in the European dataset, when focusing on the natural spline model with 8 df/year
16 (as discussed in Section 6.2.7.2) and lag 1 results (as discussed in Section 6.6.2.1),
17 respiratory mortality risk estimates were robust to the inclusion of PM10 in copollutant
18 models in all-year analyses with O3 respiratory mortality risk estimates increasing in the
19 Canadian and U.S. datasets. In summer season analyses, respiratory O3 mortality risk
20 estimates were robust in the U.S. dataset and attenuated in the European dataset.
21 Similarly, in the Italian multicity study (Stafoggia et al.. 2010). which was limited to the
22 summer season, respiratory mortality risk estimates were attenuated in copollutant
23 models with PMi0. Based on the APHENA and Italian multicity results, O3 respiratory
24 mortality risk estimates appear to be moderately to substantially sensitive (e.g., increased
25 or attenuated) to inclusion of PMi0. However, in the APHENA study, the mostly every-
26 6th-day sampling schedule for PM10 in the Canadian and U.S. datasets greatly reduced
27 their sample size and limits the interpretation of these results.
6.2.9 Summary and Causal Determination
28 The 2006 O3 AQCD concluded that there was clear, consistent evidence of a causal
29 relationship between short-term O3 exposure and respiratory effects (U.S. EPA. 2006b).
30 This conclusion was substantiated by evidence from controlled human exposure and
31 toxicological studies indicating a range of respiratory effects in response to short-term O3
32 exposure, including pulmonary function decrements, respiratory symptoms, lung
33 inflammation, increased lung permeability, and airway hyperresponsiveness.
34 Toxicological studies provided additional evidence for O3-induced impairment of host
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1 defenses. Combined, these findings from experimental studies provided support for
2 epidemiologic evidence, in which short-term O3 exposure was consistently associated
3 with decreases in lung function in populations with increased outdoor exposures, children
4 with asthma, and healthy children; increases in respiratory symptoms and asthma
5 medication use in children with asthma; and increases in respiratory-related hospital
6 admissions and asthma-related ED visits. Short-term O3 exposure also was consistently
7 associated with all-cause and cardiopulmonary mortality; however, the contribution of
8 respiratory causes to these findings was uncertain.
9 Building on the large body of evidence presented in the 2006 O3 AQCD, recent studies
10 support associations between short-term O3 exposure and respiratory effects. Controlled
11 human exposure studies continue to provide the strongest evidence for lung function
12 decrements in young healthy adults over a range of O3 concentrations. Studies previously
13 reported mean O3-induced FEVi decrements of 6-8% at 80 ppb O3 (Adams. 2006a.
14 2003a; McDonnell et al.. 1991; Horstman et al.. 1990). and new evidence additionally
15 indicates mean FEVi decrements of 6% at 70 ppb O3 (Schelegle et al. 2009) and 2-3% at
16 60 ppb O3 (Kim etal.. 2011; Brown et al.. 2008; Adams. 2006a) (Section 6.2.1.1). In
17 healthy young adults, O3-induced decrements in FEVi do not appear to depend on gender
18 (Hazucha et al.. 2003). body surface area or height (McDonnell et al.. 1997). lung size or
19 baseline FVC (Messineo and Adams. 1990). There is limited evidence that blacks may
20 experience greater O3-induced decrements in FEVi than do age-matched whites (Que et
21 al.; Seal etal.. 1993). Healthy children experience similar spirometric responses but
22 lesser symptoms from O3 exposure relative to young adults (McDonnell et al.. 1985b).
23 On average, spirometric and symptom responses to O3 exposure appear to decline with
24 increasing age beyond about 18 years of age (McDonnell et al.. 1999; Seal et al.. 1996).
25 There is also a tendency for slightly increased spirometric responses in mild asthmatics
26 and allergic rhinitics relative to healthy young adults (Torres etal.. 1996). Spirometric
27 responses in asthmatics appear to be affected by baseline lung function, i.e., responses
28 increase with disease severity (Horstman et al.. 1995).
29 Available information from controlled human exposure studies on recovery from O3
30 exposure indicates that an initial phase of recovery in healthy individuals proceeds
31 relatively rapidly, with acute spirometric and symptom responses resolving within about
32 2 to 4 h (Folinsbee and Hazucha. 1989). Small residual lung function effects are almost
33 completely resolved within 24 h. Effects of O3 on the small airways persisting a day
34 following exposure, assessed by persistent decrement in FEF25_75 and altered ventilation
35 distribution, may be due in part to inflammation (Frank etal.. 2001; Foster etal.. 1997).
36 In more responsive individuals, this recovery in lung function takes longer (as much as
37 48 hours) to return to baseline. Some cellular responses may not return to baseline levels
38 in humans for more than 10-20 days following O3 exposure (Devlin et al.. 1997). Airway
Draft - Do Not Cite or Quote 6-138 September 2011
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1 hyperresponsiveness and increased epithelial permeability are also observed as late as 24
2 h postexposure (Que etal.).
3 With repeated O3 exposures over several days, spirometric and symptom responses
4 become attenuated in both healthy individuals and asthmatics, but this tolerance is lost
5 after about a week without exposure (Gong et al.. 1997a: Folinsbee et al.. 1994; Kulle et
6 al., 1982). Airway responsiveness also appears to be somewhat attenuated with repeated
7 O3 exposures in healthy individuals, but becomes increased in individuals with
8 preexisting allergic airway disease (Gong et al.. 1997a; Folinsbee et al.. 1994). Some
9 indicators of pulmonary inflammation are attenuated with repeated O3 exposures.
10 However, other markers such as epithelial integrity and damage do not show attenuation,
11 suggesting continued tissue damage during repeated O3 exposure (Devlin et al.. 1997).
12 Collectively, epidemiologic evidence supports observations from controlled human
13 exposure studies of O3-induced decrements in lung function (Section 6.2.1.2). A notable
14 difference among newer studies was the relatively limited investigation of the effect of
15 ambient O3 exposure on lung function in populations engaged in outdoor recreation,
16 exercise, or work, which contributed to the weight of evidence in previous AQCDs. As in
17 previous AQCDs, recent epidemiologic investigation focused on and most consistently
18 demonstrated associations between increases in ambient O3 exposure and decreases in
19 lung function in children with asthma. Across the diverse populations examined in
20 epidemiologic studies, ambient O3 exposure was associated with 1-8% decreases in mean
21 lung function per standardized increment in O3 concentration1. Larger decreases (3-8%)
22 were observed in children with asthma with increased outdoor exposures, CS use, or
23 concurrent URI and older adults with airway hyperresponsiveness, elevated BMI, or
24 GSTP1 Val/Val genotype, indicating the existence of groups within the population with
25 potentially increased sensitivity to O3 exposure. Further, several epidemiologic studies
26 found that O3-associated decreases in lung function were associated with concomitant
27 increases in respiratory symptoms. Biological plausibility for O3-associated decrements
28 in lung function in controlled human exposure, epidemiologic, and animal studies is
29 provided by the well-documented effects of O3 activating bronchial C-fibers (Section
30 5.3.2).
31 Across disciplines, studies have examined factors that may potentially increase an
32 individual's susceptibility to O3-induced decrements in lung function. In the controlled
33 human exposure studies, there is a large degree of intersubject variability in lung function
34 decrements, symptomatic responses, pulmonary inflammation, airway
35 hyperresponsiveness, and altered epithelial permeability in healthy adults exposed to O3
1 Effect estimates were standardized to a 40-ppb increase for 1-h max O3, a 30-ppb increase for 8-h max O3, and a 20-ppb
increase for 24-h avg O3.
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1 (Que et al.; Holz et al.. 2005; McDonnell. 1996). The magnitude of pulmonary
2 inflammation, airway hyperresponsiveness, and increases in epithelial permeability do
3 not appear to be correlated, nor are these responses to O3 correlated with changes in lung
4 function, suggesting that different mechanisms may be responsible for these processes
5 (Que et al.; Balmes et al., 1997; Balmes et al., 1996; Aris et al., 1995). However, these
6 responses tend to be reproducible within a given individual over a period of several
7 months indicating differences in the intrinsic responsiveness of individuals (Holz et al..
8 2005: Hazuchaetal.. 2003: Holzetal.. 1999: McDonnell et al.. 1985a). Numerous
9 reasons for differences in the susceptibility of individuals to O3 exposure have been
10 reported in the literature. Dosimetric and mechanistic considerations are discussed in
11 Section 5.4. Evidence in all three disciplines suggests a role for antioxidant defenses in
12 modulating respiratory responses to O3. The biological plausibility of these findings is
13 provided by the well-characterized evidence for O3 exposure leading to the formation of
14 secondary oxidation products, which subsequently activate neural reflexes that mediate
15 lung function decrements (Section 5.3.2). Secondary oxidation products also initiate
16 pulmonary inflammation (Sections 5.3.3). Epidemiologic studies additionally have found
17 that atopy (Khatri et al., 2009). concurrent respiratory infection (Lewis et al.. 2005).
18 AHR, and elevated BMI (Alexeeff et al.. 2007) may modify respiratory responses to O3
19 exposure (Section 6.2.1.2). Retrospective analyses of controlled human exposure studies
20 of data pooled across 15 controlled human exposure studies also show larger O3-induced
21 FEVi decrements in adults with higher BMI (McDonnell et al.. 2010: Bennett et al..
22 2007).
23 Additional respiratory effects induced by short-term O3 exposures in controlled human
24 exposure studies of healthy, young adults include increases in respiratory symptoms with
25 O3 concentrations <80 ppb (Schelegle et al.. 2009: Adams. 2006a) (Section 6.2.1.1).
26 Similarly, epidemiologic studies collectively demonstrate that increases in short-term
27 ambient O3 exposure are associated with increases in respiratory symptoms and asthma
28 medication use among subjects with asthma (Section 6.2.4.1). Among recent
29 epidemiologic studies, the strongest evidence of O3-associated respiratory symptoms was
30 found in populations with multiple potential susceptibility factors, specifically,
31 individuals with asthma and atopy (Khatri et al.. 2009: Escamilla-Nunez et al.. 2008: Feo
32 Brito et al.. 2007) and children with asthma with diminished antioxidant enzyme activity
33 (Romieu et al.. 2006).
34 Recent controlled human exposure studies (Section 6.2.3.1) and toxicological studies
35 (Section 6.2.3.3) also continue to demonstrate lung injury and inflammatory responses
36 upon O3 exposure. Evidence from more than a hundred toxicological studies clearly
37 indicates that O3 induces damage and inflammation in the lung, and studies continue to
Draft - Do Not Cite or Quote 6-140 September 2011
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1 elucidate the mechanistic pathways involved (Section 5.3). Though inflammation may
2 resolve, continued inflammation may alter theJanetrita2010
3 structure and function of pulmonary tissues. New controlled human studies support
4 previous findings for pulmonary inflammation at 60 ppb O3, the lowest concentration
5 evaluated . Building on the extensive experimental evidence, epidemiologic studies
6 provide new evidence for ambient O3-associated increases in pulmonary inflammation in
7 individuals with asthma. These associations were observed primarily for 8-h max or 8-h
8 average O3 exposures but for both same-day and multiday average exposures. Multiple
9 studies examined and found increases in eNO (Berhane et al.. 2011; Khatri et al.. 2009;
10 Barraza-Villarreal et al., 2008). The clinical significance of these findings was supported
11 by observations of concomitant O3-associated increases in respiratory symptoms (Khatri
12 et al., 2009; Barraza-Villarreal et al.. 2008). A smaller number of studies examined and
13 found associations with cytokines such as IL-6 or IL-8 in nasal lavage samples (Barraza-
14 Villarreal et al.. 2008; Sienra-Monge et al., 2004) inflammatory cells in blood (e.g.,
15 eosinophils) (Khatri et al.. 2009). decreased levels of antioxidants (Sienra-Monge et al..
16 2004). and increased levels of indicators of oxidative stress (Romieu et al.. 2008)
17 (Section 6.2.3.2).
18 Modification of innate and adaptive immunity is emerging as a mechanistic pathway
19 underlying the effects of ozone on asthma and allergic airways disease (Section 5.3.6).
20 While the majority of evidence comes from animal studies, results from controlled
21 human exposure studies suggest that these pathways may be relevant to humans and may
22 lead to the induction and exacerbation of asthma (Alexis et al.. 2010; Hernandez et al..
23 2010; Alexis et al.. 2009; Bosson et al.. 2003). Further, differences between asthmatics
24 and healthy controls in ozone-mediated innate and adaptive immune responses have been
25 noted (Section 5.4.2.2).
26 The subclinical and overt respiratory effects observed across disciplines collectively
27 provide support for epidemiologic studies that demonstrate consistently positive
28 associations of short-term O3 exposure with respiratory-related hospital admissions and
29 ED visits (Section 6.2.7). Consistent with evidence presented in the 2006 O3 AQCD, new
30 multicity studies and a multicontinent study (i.e., APHENA) (Katsouyanni et al.. 2009)
31 found risk estimates ranging from an approximate 1.6 to 5.4% increase in all respiratory-
32 related hospital admissions and ED visits in all-year analyses for standardized increases
33 in ambient O3 concentrations1. Positive associations persisted in analyses restricted to the
34 summer season, but the magnitude varied depending on the study location (Figure 6-19).
35 Compared with studies reviewed in the 2006 O3 AQCD, a larger number of recent studies
1 Effect estimates were standardized to a 20-ppb increase for 24-h avg O3, a 30-ppb increase for 8-h max O3, and a 40-ppb
increase for 1-h max O3.
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1 examined hospital admissions and ED visits for specific respiratory outcomes. Although
2 still limited in number, both single- and multicity studies found consistent, positive
3 associations between short-term O3 exposures and asthma and COPD hospital admissions
4 and ED visits, with more limited evidence for pneumonia. Consistent with the
5 conclusions of the 2006 O3 AQCD, in studies that conducted seasonal analyses, risk
6 estimates were elevated in the warm season compared to cold season or all-season
7 analyses, specifically for asthma and COPD. Although recent studies did not include
8 detailed age-stratified results, the increased risk of asthma hospital admissions
9 (Silverman and Ito, 2010; Strickland et al., 2010; Dales et al., 2006) observed for children
10 strengthens the conclusion from the 2006 O3 AQCD that children are particularly
11 susceptible to O3-induced respiratory effects (U.S. EPA. 2006b). Although the
12 concentration-response relationship has not been extensively examined, preliminary
13 examinations found no evidence of a threshold between short-term O3 exposure and
14 asthma hospital admissions and pediatric asthma ED visits (Silverman and Ito. 2010;
15 Strickland etaL 2010).
16 New evidence extends the potential range of well-established O3-associated respiratory
17 effects by demonstrating associations between short-term ambient O3 exposure and
18 respiratory-related mortality. In all-year analyses, a multicontinent (APHENA) and
19 multicity (PAPA) study found consistent, positive associations with respiratory mortality
20 for all ages but less consistent evidence in analyses restricted to ages 75+. Further,
21 multicity studies in the U.S. and Europe that conducted seasonal analyses found stronger
22 associations during the summer season (Section 6.2.8).
23 Several studies of respiratory morbidity and mortality evaluated the potential
24 confounding effects of copollutants, in particular, PM10, PM2 5, or NO2. In most cases,
25 effect estimates remained robust to the inclusion of copollutants; however, in several
26 studies, changes were observed in the magnitude of the O3 association. In studies of lung
27 function and respiratory symptoms, larger effects frequently were estimated for O3 when
28 copollutants were added to models. Ozone effect estimates for respiratory-related hospital
29 admissions and ED visits remained relatively robust upon the inclusion of PM and
30 gaseous pollutants in two-pollutant models (Strickland et al., 2010; Tolbert et al., 2007;
31 Medina-Ramon et al. 2006). Although copollutant confounding was not extensively
32 examined in mortality studies, the O3-respiratory mortality relationship was moderately
33 to substantially sensitive (e.g., increased or attenuated) to the inclusion of PMi0 in
34 copollutant models (Stafoggia et al., 2010; Katsouvanni et al.. 2009). However,
35 interpretation of these results requires caution due to the limited PM datasets used in
36 these studies. Together, these findings across respiratory endpoints provide support for
37 the independent effects of short-term ambient O3 exposures.
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1 In summary, new studies evaluated since the completion of the 2006 O3 AQCD support
2 and expand upon the strong body of evidence that indicated a causal relationship between
3 short-term O3 exposure and respiratory health effects. New controlled human exposure
4 studies continue to demonstrate O3-induced decreases in FEVi and pulmonary
5 inflammation at concentrations as low as 60 ppb. New epidemiologic studies provide
6 evidence for associations of ambient O3 exposure with biological markers of pulmonary
7 inflammation and oxidative stress. Toxicological studies have continued to support the
8 biological plausibility for the O3-induced respiratory effects observed in the controlled
9 human exposure and epidemiologic studies. Additionally, recent epidemiologic studies
10 further confirm that respiratory morbidity and mortality associations are stronger during
11 the warm/summer months and remain relatively robust after adjustment for copollutants.
12 However, despite the consistency of association between short-term O3 exposure and
13 respiratory effects, new evidence suggests that the magnitude of association may be
14 underestimated due to behavioral modification in response to forecasted air quality
15 (Section 4.6.4). Collectively, the new evidence integrated across toxicological, controlled
16 human exposure, and epidemiologic studies, in conjunction with that reviewed in
17 previous AQCDs, is sufficient to conclude that there is a causal relationship between
18 short-term O3 exposure and respiratory health effects.
6.3 Cardiovascular Effects
6.3.1 Controlled Human Exposure
19 O3 reacts rapidly on contact with respiratory system tissue and is not absorbed or
20 transported to extrapulmonary sites to any significant degree as such. Controlled human
21 exposure studies discussed in the previous AQCDs failed to demonstrate any consistent
22 extrapulmonary effects. Some controlled human exposure studies have attempted to
23 identify specific markers of exposure to O3 in blood. Foster et al. (1996) found a
24 reduction in the serum levels of the free radical scavenger a-tocopherol after O3 exposure.
25 Liu et al. (1999; 1997) used a salicylate metabolite, 2,3, dehydroxybenzoic acid (DHBA),
26 to indicate increased levels of hydroxyl radical which hydroxylates salicylate to DHBA.
27 Increased DHBA levels after exposure to 120 and 400 ppb suggest that O3 increases
28 production of hydroxyl radical. The levels of DHBA were correlated with changes in
29 spirometry.
30 Gong et al. (1998) observed a small, statistically significant O3-induced increase in the
31 alveolar-to-arterial PO2 gradient in both healthy (n = 6) and hypertensive (n = 10) adult
32 males (aged 41-78 years) exposed for 3 hours with exercise to 300 ppb O3. The
Draft - Do Not Cite or Quote 6-143 September 2011
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1 mechanism for the decrease in arterial oxygen tension in the Gong et al. (1998) study
JO O V / J
2 could be due to an O3-induced ventilation-perfusion mismatch. Gong et al. (1998)
3 suggested that by impairing alveolar-arterial oxygen transfer, the O3 exposure could
4 potentially lead to adverse cardiac events by decreasing oxygen supply to the
5 myocardium. The subjects in the Gong et al. (1998) study had sufficient functional
6 reserve so as to not experience significant ECG changes or myocardial ischemia and/or
7 injury. In studies evaluating the exercise performance of healthy adults, no significant
8 effect of O3 on arterial O2 saturation has been observed (Schelegle and Adams. 1986).
9 More recently, Fakhri et al. (2009) evaluated changes in HRV among adult volunteers
10 (n=50; 27 ± 7 years) during 2-h exposures to PM2 5 CAPs (127±62 ug/m3) and O3
11 (114±7 ppb), alone and in combination. High frequency HRV was increased following
12 CAPs-only (p=0.046) and O3-only (p=0.051) exposures, but not in combination. The
13 standard deviation of NN intervals and the square root of the mean squared differences of
14 successive NN intervals also showed marginally significant (0.05
-------
6.3.2.1 Arrhythmia
1
2
3
4
5
6
7
8
9
10
11
12
In the 2006 O3 AQCD, conflicting results were observed when examining the effect of O3
on arrhythmias (Dockery et al.. 2005; Rich et al., 2005). A study by Dockery et al. (2005)
reported no association between O3 levels and ventricular arrhythmias among patients
with implantable cardioverter defibrillators (ICD) living in Boston, MA, although when
O3 was categorized into quintiles, there was weak evidence of an association with
increasing O3 concentration (median O3 concentration: 22.9 ppb). Rich et al. (2005)
performed a re-analysis of this cohort using a case-crossover design and detected a
positive association between O3 exposure and ventricular arrhythmias. Recent studies
were conducted in various locations and each used a different cardiac episode to define
an arrhythmic event and a different time period of exposure, which may help explain
observed differences across studies. Ozone levels for each new study are reported in
Table 6-29.
13
14
15
16
17
18
19
20
21
Table 6-29 Characterization of ozone concentrations (in ppb) from studies of
arrhythmias
'Reference Location
Anderson et al. (2010) London, England
Metzger et al. (2007) Atlanta, GA
Rich et al. (2006a) St. Louis, MO
Rich et al. (2006b) Boston, MA
Sarnat et al. (2006a) Steubenville, OH
Averaging Time
8-h max
8-h max
Summer only
24-h
1-h
24-h
24-h
Summer and Fall only
5 days
Mean Concentration
(Standard Deviation)
8.08
53.9 (23)
21*
22.2*
22.6*
21.8(12.6)
22.2(9.1)
Upper Range
of Concentration
75th: 11. 5
Max: 148
75th: 31
75th: 33
Max: 119.5
75th: 30.9
Max: 77.5
75th: 28.5
Max: 74.8
75th: 29.1
Max: 44
'Median presented (information on mean not given).
Multiple studies examined O3-related effects on individuals with ICDs. One study of 518
ICD patients who had at least 1 tachyarrythmia within a 10-year period (totaling 6287
tachyarrhythmic event-days; 1993-2002) was conducted in Atlanta, Georgia (Metzger et
al.. 2007). Tachyarrhythmic events were defined as any ventricular tachyarrhythmic
event, any ventricular tachyarrhythmic event that resulted in electrical therapy, and any
ventricular tachyarrhythmic event that resulted in defibrillation. In the primary analysis,
no evidence of an association was observed for a 30 ppb increase in 8-h max O3
concentrations and tachyarrhythmic events (OR: 1.00 [95% CI: 0.92, 1.08]; lag 0).
Season-specific as well as several sensitivity analyses (including the use of an
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1 unconstrained distributed lag model [lags 0-6]) were conducted resulting in similar null
2 associations. A strength of this study is that it incorporated a large sample size over a
3 long time period.
4 In a case-crossover analysis, a population of ICD patients in Boston, previously examined
5 by (Rich et al.. 2005) was used to assess the association between air pollution and
6 paroxysmal atrial fibrillation (PAF) episodes (Rich et al.. 2006b). In addition to
7 ventricular arrhythmias, ICD devices may also detect supraventricular arrhythmias, of
8 which atrial fibrillation is the most common. Although atrial fibrillation is generally not
9 considered lethal, it has been associated with increased premature mortality as well as
10 hospitalization and stroke. Ninety-one electrophysiologist-confirmed episodes of PAF
11 were ascertained among 29 patients. An association (OR: 3.86 [95% CI: 1.44, 10.28] per
12 40 ppb increase in 1-h max O3 concentrations) was observed between increases in O3
13 during the concurrent hour and PAF episodes (lag 0-h). The estimated OR for the 24-h
14 moving average concentration was elevated (OR: 1.81 [95% CI: 0.86, 3.83] per 20 ppb),
15 but weaker than the estimate for the shorter exposure window. The association between
16 PAF and O3 in the concurrent hour during the cold months was comparable to that during
17 the warm months. In addition, no evidence of a deviation from linearity between O3
18 concentration and the log odds of PAF was observed. Authors report that the difference
19 between O3 exposure and observed effect between this study (PAF and 1-h O3) and their
20 previous study (ventricular arrhythmias and 24-h moving average O3) (Rich et al.. 2005)
21 suggest a more rapid response to air pollution for PAF (Rich et al., 2006b).
22 In an additional study, Rich et al. (2006a) employed a case-crossover design to examine
23 the association between air pollution and 139 confirmed ventricular arrhythmias among
24 56 ICD patients in St Louis, Missouri. The authors observed a positive association with
25 O3 (OR: 1.17 [95% CI: 0.58, 2.38] per 20 ppb increase in 24-h moving avg O3
26 concentrations [lags 0-23 hours]). Although the authors concluded these results were
27 similar to their results from Boston (Rich et al.. 2005). they postulated that the pollutants
28 responsible for the increased risk in ventricular arrhythmias are different (O3 and PM2 5 in
29 Boston and sulfur dioxide in St Louis).
30 Anderson et al. (2010) used a case-crossover framework to assess air pollution and
31 activation of ICDs among patients from all 9 ICD clinics in the London National Health
32 Service hospitals. "Activation" was defined as tachycardias for which the defibrillator
33 delivered treatment. Investigators modeled associations using unconstrained distributed
34 lags from 0 to 5 days. The sample consisted of 705 patients with 5,462 activation days
35 (O3 information was for 543 patients and 4,092 activation days). Estimates for O3 were
36 consistently positive, although weak (OR: 1.09 [95% CI: 0.76, 1.55] per 30 ppb for 0-
Draft - Do Not Cite or Quote 6-146 September 2011
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1 1 day lag; OR: 1.04 [95% CI: 0.60, 1.81] per 30 ppb for 0-5 day lag) (Anderson et al.
2 2010).
3 In contrast to arrhythmia studies conducted among ICD patients, Sarnat et al. (2006a)
4 recruited non-smoking adults (age range: 54-90 years) to participate in a study of air
5 pollution and arrhythmias conducted over two 12-week periods during summer and fall
6 of 2000 in a region characterized by industrial pollution (Steubenville, Ohio). Continuous
7 ECG data acquired on a weekly basis over a 30-minute sampling period were used to
8 assess ectopy, defined as extra cardiac depolarizations within the atria (supraventricular
9 ectopy, SVE) or the ventricles (ventricular ectopy, VE). Increases in the 5-day moving
10 average (days 1-5) of O3 were associated with an increased odds of SVE (OR: 2.17 [95%
11 CI: 0.93, 5.07] per 20 ppb increase in 24-h avg O3 concentrations). A weaker association
12 was observed for VE (OR: 1.62 [95% CI: 0.54, 4.90] per 20 ppb increase in 24-h avg O3
13 concentrations). The results of the effect of 5-day O3 on SVE were robust to the inclusion
14 of SO42" in the model [OR: 1.62 (95% CI: 0.54, 4.90)]. The authors indicate that the
15 strong associations observed at the 5-day moving averages, as compared to shorter time
16 periods, suggests a relatively long-acting mechanistic pathways, such as inflammation,
17 may have promoted the ectopic beats in this population (Sarnat et al.. 2006a).
18 Although many studies report positive associations, collectively, studies of arrhythmias
19 report inconsistent results. This may be due to variation in study populations, length and
20 season of averaging time, and outcome under study. Future studies are expected to
21 provide additional evidence for the various outcomes and exposure periods.
6.3.2.2 Heart Rate/Heart Rate Variability
22 In the 2006 O3 AQCD, two large population-based studies of air pollution and HRV were
23 summarized (Park et al.. 2005b: Liao et al.. 2004a). In addition, the biological
24 mechanisms and potential importance of HRV were discussed. Briefly, the study of acute
25 adverse effects of air pollution on cardiac autonomic control is based on the hypothesis
26 that increased air pollution levels may stimulate the autonomic nervous system and lead
27 to an imbalance of cardiac autonomic control characterized by sympathetic activation
28 unopposed by parasympathetic control (U.S. EPA. 2006b). Examples of HRV indices
29 include the standard deviation of normal-to-normal intervals (SDNN), the square root of
30 the mean of the sum of the squares of differences between adjacent NN intervals (r-
31 MSSD), high-frequency power (HF), low-frequency power (LF), and the LF/HF ratio.
32 Liao et al. (2004a) examined the association between air pollution and cardiac autonomic
33 control in the fourth cohort examination (1996-1998) of the U.S.-based Atherosclerosis
34 Risk in Communities Study. A decrease in log-transformed HF was associated with an
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
increase in O3 concentration among white study participants. Park et al. (2005b)
examined the effects of air pollution on indices of HRV in a population-based study
among men from the Normative Aging Study in Boston, Massachusetts. Several
associations were observed with O3 and HRV outcomes; a reduction in LF was associated
with increased O3 concentration, which was robust to inclusion of PM25. The associations
with all HRV indices and O3 were stronger among those with ischemic heart disease and
hypertension. In addition to these population-based studies included in the 2006 O3
AQCD was a study by Schwartz et al. (2005). who conducted a panel study to assess the
relationship between exposure to summertime air pollution and HRV. A weak association
of O3 during the hour immediately preceding the health measures was observed with r-
MSSD among a study population that consisted of mostly older female participants. In
summary, these studies suggest that short-term exposures to O3 are predictors of
decreased HRV and that the relationship may be stronger among certain subgroups. The
generally consistent (although weak) associations between pollutants and reduced cardiac
autonomic control were observed at relatively low pollution concentrations typically
recent studies of O3 and
studies are presented in
HRV and are described below. The O3 concentrations for these
Table 6-30.
Table 6-30 Characterization of ozone concentrations (in ppb) from studies of
heart rate variability
Reference Location
Baja et al. (2010) Boston, MA
Chan et al. (2005a) Taipei, Taiwan
Chuang et al. (2007a) Taipei, Taiwan
Chuang et al. (2007b) Taipei, Taiwan
Park etal. (2007) Boston, MA
Park etal. (2008) Boston, MA
Ruidavets et al. (2005a) Toulouse, France
Wheeler et al. (2006) Atlanta, GA
Wu et al. (2010) Taipei, Taiwan
Zanobetti et al. (2010) Boston, MA
Averaging Time
Olag
10-hlag
1-h
24-h
48-h
72-h
1-h
24-h
24-h
8-h
4-h
24-h
Working period
0.5-h
2-h
3-D
5-D
Mean Concentration
(Standard Deviation)
23(16)
21 (15)
21.9(15.4)
28.4(12.1)
33.3 (8.9)
33.8(7.1)
35.1
Range of 17.0-29.1
23.4(13)
38.3(14.8)
18.5
29.4
24.9(14.0)
20.7*
20.5*
21.9*
22.8*
Upper Range of
Concentration
Max: 114.9
Max: 49.3
Max: 47.8
Max: 48.3
Max: 192.0
75th: 46.9
Max: 80.3
75th: 22.5
Max: 59.2
75th: 30.33
75th: 30.08
75th: 28.33
75th: 29.28
'Median presented (information on mean not given).
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1 Several follow-up examinations of HRV were conducted among the participants of the
2 Normative Aging Study in Boston. A trajectory cluster analysis was used to assess
3 whether pollution originating from different locations had varying relationships with
4 HRV (Park et al.. 2007). Subjects who were examined on days when air parcels
5 originated in the west had the strongest associations with O3; however, the O3
6 concentration in this cluster was low (24-h avg, 17.0 ppb) compared to the other clusters
7 (24-h avg of 21.3-29.1 ppb). LF and SDNN decreased with increases in the 4-h moving
8 average of O3 from the west (LF decreased by 51.2% [95% CI: 1.6, 75.9%] and SDNN
9 decreased by 28.2% [95% CI: -0.5, 48.7%] per 30 ppb increase in 4-h avg O3
10 concentrations) (Parketal.. 2007). The Boston air mass originating in the west traveled
11 over Illinois, Indiana, and Ohio; states typically characterized by coal-burning power
12 plants. Due to the low O3 concentrations observed in the west cluster, the authors
13 hypothesize that O3 on those days could be capturing the effects of other, secondary
14 and/or transported pollutants from the coal belt or that the relationship between ambient
15 O3 and personal exposure to O3 is stronger during that period (supported by a
16 comparatively low apparent temperature which could indicate a likelihood to keep
17 windows open and reduced air conditioning use) (Park et al.. 2007). An additional
18 follow-up evaluation using the Normative Aging Study examined the potential for effect
19 modification by chronic lead exposure on the relationship between air pollution and HRV
20 (Park et al.. 2008). Authors observed graded reductions in HF and LF of HRV in relation
21 to O3 (and sulfate) across increasing quartiles of tibia and patella lead (HF: %change 32.3
22 [95% CI: -32.5, 159.3] for the first quartile of tibia Pb and -59.1 [95% CI: -77.3, -26.1]
23 for the fourth quartile of tibia Pb per 30 ppb increase in 4-h avg O3 concentrations; LF:
24 %change 8.0 [95% CI: -36.9, 84.9] for the first quartile of tibia Pb and -59.3 [95% CI: -
25 74.6, -34.8] for the fourth quartile of tibia Pb per 30 ppb increase in 4-h avg O3
26 concentrations). In addition, O3 associations were similar when education and cumulative
27 traffic-adjusted bone lead levels were used in analyses. Authors indicate the possibility
28 that O3 (which has low indoor concentrations) was acting as a proxy for sulfate
29 (correlation coefficient for O3 and sulfate = 0.57). Investigators of a more recent follow-
30 up to the Normative Aging Study hypothesized that the relationships between short-term
31 air pollution exposures and ventricular repolarization, as measured by changes in the
32 heart-rate corrected QT interval (QTc), would be modified by participant characteristics
33 (e.g., obesity, diabetes, smoking history) and genetic susceptibility to oxidative stress
34 (Bajaetal.. 2010). No evidence of an association between O3 (using a quadratic
35 constrained distributed lag model and hourly exposure lag models over a 10-h time
36 window preceding the visit) and QTc was reported (change in mean QTc -0.74 [95% CI:
37 -3.73, 2.25]); therefore, potential effect modification of personal and genetic
38 characteristics with O3 was not assessed (BajaetaL 2010). Collectively, the results from
39 studies that examined the Normative Aging Study cohort found an association between
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1 increases in short-term exposures to O3 and decreases in HRV (Park et al., 2008; Park et
2 al.. 2007; Park et al.. 2005b) although not consistently in all of the studies (Baja et al..
3 2010). Further, observed relationships appear to be stronger among those with ischemic
4 heart disease, hypertension, and elevated bone lead levels, as well as when air masses
5 arrive from the west (the coal belt). However, it is not clear if O3 is acting as a proxy for
6 other, secondary particle pollutants (such as sulfate) (2008; 2007; Parketal.. 2005b). In
7 addition, since the Normative Aging Study participants were older, predominately white
8 men, results may not be generalizable to women, younger individuals, or those of
9 different racial/ethnic groups (Bajaet al., 2010).
10 Additional studies of populations not limited to the Normative Aging Study have also
11 examined associations between O3 exposure and HRV. A panel study among 18
12 individuals with COPD and 12 individuals with recent myocardial infarction (MI) was
13 conducted in Atlanta, Georgia (Wheeler et al.. 2006). HRV was assessed for each
14 participant on 7 days in fall 1999 and/or spring 2000. The mean 4-h O3 concentration
15 (time period immediately preceding the HRV measures) was 18.5 ppb; however, O3
16 concentrations differed substantially within study sites (8.0 - 33.8 ppb). Ozone
17 concentrations were not associated with HRV (SDNN) among all subjects (percent
18 change of 2.36% [95% CI: -10.8%, 17.5%] per 30 ppb 4-h O3 increase) or when stratified
19 by disease type (COPD, recent MI, and baseline FEVO (Wheeler et al.. 2006).
20 HRV and air pollution was assessed in a panel study among 46 predominately white male
21 patients (study population: 80.4% male, 93.5% white) aged 43-75 years in Boston,
22 Massachusetts, with coronary artery disease (Zanobetti et al., 2010). Up to four home
23 visits were made to assess HRV over the year following the index event. Pollution lags
24 used in analyses ranged between 30 minutes to a few hours and up to 5 days prior to the
25 HRV assessments. Decreases in r-MSSD were reported for all averaging times of O3
26 (percent change of-5.18% [95% CI: -7.89, -2.30] per 20 ppb of 5-day moving average of
27 O3 concentration), but no evidence of an association between O3 and HF was observed
28 (quantitative results not provided). In two-pollutant models with O3 and either PM2 5 or
29 BC, O3 associations remained robust.
30 A few studies were conducted outside of the U.S. to assess the relationship between air
31 pollution concentrations and heart rate and HRV (Wu et al., 2010; Chuang et al., 2007b;
32 Chuang et al.. 2007a; Chan et al.. 2005a; Ruidavets et al.. 2005a). No associations were
33 reported between O3 and HRV among CHD patients and patients with one or more major
34 CHD risk factors residing in Taipei, Taiwan (Chan et al.. 2005a). Another study in
3 5 Taipei, Taiwan examined mail carriers and reported O3 levels measured using personal
36 monitors. No association was observed between O3 and the measures of HRV (percent
37 change for SDNN: 0.57 [95% CI: -21.27, 28.46], r-MSSD: -7.10 [95% CI: -24.24, 13.92],
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1 HF: -1.92 [95% CI: -23.68, 26.02], LF: -4.82 [95% CI: -25.34, 21.35] per 40 ppb O3)
2 ("Wuetal.. 2010). In addition, no consistent relationships were identified between O3 and
3 resting heart rate among middle-aged (35-64 years) participants residing in Toulouse,
4 France (Ruidavets et al.. 2005a). A negative trend was reported for the 3-day cumulative
5 (lag days 1-3) concentration of O3 with heart rate (p for trend = 0.02); however, the
6 individual odds ratios comparing quintiles of exposure showed no association (OR for O3
7 ofO.93 [95% CI: 0.86, 1.01] for the highest quintile of resting heart rate compared to the
8 lowest). When stratified by current smoking status, non-smokers had a decreased trend
9 with increased 3-day cumulative O3 concentrations but none of the quintiles for heart rate
10 were statistically significant. A panel study was conducted in Taiwan to assess the
11 relationship between air pollutants and inflammation, oxidative stress, blood coagulation,
12 and autonomic dysfunction (Chuang et al.. 2007b: Chuang et al.. 2007a). Participants
13 were apparently healthy college students (aged 18-25 year) who were living in a
14 university dormitory in metropolitan Taipei. Health endpoints were measured three times
15 from April to June in 2004 or 2005. Ozone was assessed in statistical models using the
16 average of the 24, 48, and 72 hours before the hour of each blood sampling. Decreases in
17 HRV (measured as SDNN, r-MSSD, LF, and HF) were associated with increases in O3
18 concentrations in single-pollutant models (percent change for SDNN: -13.45 [95% CI: -
19 16.26, -10.60], r-MSSD -13.76 [95% CI: -21.62, -5.44], LF -9.16 [95% CI: -13.29, -
20 4.95], HF -10.76 [95% CI: -18.88, -2.32] per 20 ppb 3-day avg O3 concentrations) and
21 remained associated with 3-day O3 concentrations in two-pollutant models with sulfate.
22 Another study in Taiwan recruited individuals with coronary heart disease or at risk for
23 cardiovascular disease from outpatient clinics (Chuang et al.. 2007b). Mean O3
24 concentrations were 35.1 ppb (SD 27.5 ppb) during the study period (two weeks in
25 February). No association was observed between O3 concentration and HRV measures
26 (SDNN, r-MSSD, LF, HF) (numerical results not provided in publication).
27 Overall, studies of O3 concentration and HRV report inconsistent results. Multiple studies
28 in Boston observed positive associations but the authors of many of these studies
29 postulated that O3 was possibly acting as a proxy for other pollutants. The majority of
30 other studies, both in the U.S. and internationally, report null findings. The
31 inconsistencies observed are further complicated by the different HRV measures and
32 averaging times used by the studies.
6.3.2.3 Stroke
33 The 2006 O3 AQCD did not identify any studies that examined the association between
34 short-term O3 exposure and stroke. However, recent studies have attempted to examine
35 this relationship. Lisabeth et al. (2008) used a time-series approach to assess the
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1 relationship between daily counts of ischemic stroke and transient ischemic attack (TIA)
2 with O3 concentrations in a southeast Texas community among residents 45 years and
3 older (2001-2005; median age of cases, 72 years). The median O3 (hourly average per 24-
4 h time-period) concentration was 25.6 ppb (IQR 18.1-33.8). The associations between
5 same-day (RR: 1.03 [95% CI: 0.96, 1.10] per 20 ppb increase in 24-h avg O3
6 concentrations) and previous-day (RR: 1.05 [95% CI: 0.99, 1.12] per 20 ppb increase in
7 24-h avg O3 concentrations) O3 concentrations and stroke/TIA risk were positive.
8 Associations were robust to adjustment for PM2 5. The effect of season on the relationship
9 was not assessed.
10 A case-crossover design was used in a study conducted in Dijon, France between March
11 1994 and December 2004, among those 40 years of age and older who presented with
12 first-ever stroke (Henrotin et al., 2007). The mean O3 concentration, calculated over 8-h
13 daytime periods, was 14.95 ppb (IQR: 6-22 ppb). No association was observed between
14 O3 concentration at 0, 1, 2, or 3 days lag and hemorrhagic stroke. However, an
15 association between ischemic stroke occurrence and O3 concentrations with a 1-day lag
16 was observed (OR: 1.54 [95% CI: 1.14, 2.09] per 30 ppb increase in 8-h max O3
17 concentrations). The effect of O3 persisted in two-pollutant models with PMi0, SO2, NO2,
18 or CO. This association was stronger among men (OR: 2.12 [95% CI: 1.36, 3.30] per 30
19 ppb increase in 8-h max O3 concentrations) than among women (OR: 1.17 [95%CI: 0.77,
20 1.78] per 30 ppb increase in 8-h max O3 concentrations) in single pollutant models. When
21 stroke was examined by subtype among men, an association was observed for ischemic
22 strokes of large arteries and for transient ischemic attacks but not for cardioembolic or
23 lacunar ischemic strokes. The subtype analysis was not performed for women.
24 Additionally, for men a linear exposure-response was observed when O3 was assessed
25 based on quintiles (p for trend = 0.01) (Figure 6-21). A potential limitation of this study
26 is that 67.4% of the participating men were smokers compared to 9.3% of the women.
27 Another study, performed in Dijon, France, examined the association between O3
28 concentration and incidence of fatal and non-fatal ischemic cerebrovascular events
29 (ICVE) (Henrotin etal.. 2010V Mean 8-h O3 concentration was 19.1 ppb (SD 12.2 ppb).
30 A positive association was observed between recurrent ICVE and O3 concentration with a
31 3-day lag (OR: 1.92 [95% CI 1.17, 3.12]), butnotfor other lags (0, 1, 2, 4) or cumulative
32 days (0-1, 0-2, 1-2, 1-3). Although some ORs for incident ICVEs were elevated, none
33 were statistically significant. Results for associations using the maximum daily 1-h O3
34 concentrations were similar to the 8-h results but slightly attenuated. ORs were similar in
35 two pollutant models (data not given). In stratified analyses, the association between 1-
36 day lagged O3 concentration and incident and recurrent ICVE was greater among those
37 with multiple other preexisting vascular conditions. Increased associations with ICVE
38 were also observed for individuals with diabetes or hypertension.
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3.5
3 -
2.5 -
o
I 2
1.5 -
0.5 -
0-8 9-20 21-32 33-48
O3 concentration (ppb)
Source: Henrotin et al. (2007).
Figure 6-21 Odds ratio (95% confidence interval) for stroke by quintiles of
ozone
1
2
3
4
5
6
6.3.2.4 Biomarkers
An increasing number of studies have examined the relationship between air pollution
and biomarkers in an attempt to elucidate the biological mechanisms linking air pollution
and cardiovascular disease. A wide range of markers assessed as well as different types
of study designs and locations chosen make comparisons across studies difficult. Table 6-
31 provides an overview of the O3 concentrations reported in each of the studies
evaluated.
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Table 6-31 Characterization of ozone concentrations (in ppb) from studies of
biomarkers
Reference
Baccarelli et al. (2007)
Chen etal. (2007)
Chuanaetal. (2010)
Chuang et al. (2007a)
Goldberg et al. (2008)
Liao et al. (2005)
Rudez etal. (2009)
Steinvil etal. (2008)
Thompson etal. (2010)
Wellenius et al. (2007)
Location
Lombardia, Italy
Los Angeles and
San Francisco, CA
Taiwan
Taipei, Taiwan
Montreal, Quebec
3 U.S. counties
Rotterdam, the Netherlands
Tel -Aviv, Israel
Toronto, Ontario
Boston, MA
Averaging
Time
1-h
8-h/2 wk
8-h/1 mo
24-h
48-h
72-h
24-h
8-h
24-h
0.5-h
1-h/1yr
1-h/24-h
Mean Concentration
(Standard Deviation)
18.3*
30.8*
28.3*
26.83 (9.7)
28.4(12.1)
33.3 (8.9)
33.8(7.1)
NS
40 (20)
22*
29.2 (9.7)
21.94(15.78)
25.1 (12.9)
Upper Range of
Concentration
75th: 35.1
Max: 202.3
Max: 47.9
Max: 43.1
Max: 62.1
Max: 49.3
Max: 47.8
Max: 48.3
75th: 31 .5
Max: 90
75th: 36
'Median presented (information on mean not given).
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
Hemostasis and coagulation markers
Multiple studies used various markers to examine if associations were present between
O3 concentrations and hemostatis and coagulation. Some of the markers included in these
studies were as follows: fibrinogen, von Willebrand factor (vWF), plasminogen activator
fibrinogen inhibitor-1 (PAI-1), tissue-type plasminogen activator (tPA), platelet
aggregation, and thrombin generation.
A population-based study in the United States was conducted to assess the relationship
between short-term exposure to air pollution and markers of blood coagulation using the
Atherosclerosis Risk in Communities (ARIC) study cohort (Liao etal.. 2005). Significant
curvilinear associations were observed for O3 (1 day prior to blood draw) and fibrinogen
and vWF (quantitative results not provided for regression models although adjusted
means [SE] of vWF were given as 118% [0.79%] for O3 concentrations <40 ppb, 117%
[0.86%] for O3 concentrations 40-70 ppb, and 124% [1.97%] for O3 concentrations of 70
ppb). The association between O3 and fibrinogen was more pronounced among those
with a history of cardiovascular disease (CVD) and was statistically significant among
only this subgroup of the population. The curvilinear relationship between exposure and
outcome suggested stronger relationships at higher concentrations of O3 which could
indicate threshold effects. The authors note that the most pronounced associations
occurred when the pollutants were 2-3 standard deviations above the mean. The results
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1 from this relatively large-scale cross-sectional study suggest weak associations with O3
2 and fibrinogen (among those with a history of CVD) and vWF.
3 A retrospective repeated measures analysis was performed in Toronto, Canada among
4 adults aged 18-40 years (n=45) between the years of 1999 and 2006 (Thompson et al.,
5 2010). Single pollutant models were used with moving averages up to 7 days. No
6 evidence of an association was observed for O3 and fibrinogen.
7 A repeated measures study was conducted in 40 healthy individuals living or working in
8 the city center of Rotterdam, the Netherlands to assess the relationship between air
9 pollution and markers of hemostatis and coagulation (platelet aggregation, thrombin
10 generation, and fibrinogen) (Rudez et al.. 2009). Each participant provided between 11
11 and 13 blood samples throughout a 1-year period (498 samples on 197 days). Examined
12 lags ranged from 6 hours to 3 days prior to blood sampling. No consistent evidence of an
13 association was observed between O3 and any of the biomarkers (percent change of max
14 platelet aggregation: -6.87 [95% CI: -21.46, 7.70] per 20 ppb 4-day average O3; percent
15 change of endogenous thrombin potential: 0.95 [95% CI: -3.05, 4.95] per 20 ppb 4-day
16 avg O3; percent change of fibrinogen: -0.57 [95% CI: -3.05, 2.00] per 20 ppb lag 1-day
17 O3). Some associations with O3 were in the opposite direction to that hypothesized which
18 may be explained by the negative correlation between O3 and the other pollutants
19 (correlation coefficients ranged from -0.4 to -0.6). The statistically significant inverse
20 effects observed with O3 in single-pollutant models were no longer apparent when PM10
21 was included in the models (Rudez et al.. 2009).
22 A panel study in Taiwan measured health endpoints using blood samples from healthy
23 individuals (n=76) at three times from April to June in 2004 or 2005 (Chuang et al..
24 2007a). Increases in fibrinogen and PAI-1 were associated with increases in O3
25 concentrations in single-pollutant models (percent change in fibrinogen: 11.76 [95% CI:
26 4.03, 19.71] per 20 ppb 3-day avg O3; percent change in PAI-1: 6.08 [95% CI: 38.91,
27 84.27] per 20 ppb 3-day avg O3). These associations were also observed at 1 and 2 day
28 averaging times. Associations between PAI-1 and 3-day O3 concentrations remained
29 robust in two-pollutant models with sulfate. No association was seen between O3 and
30 tPA, a fibrinolytic factor (percent change 16.15 [95% CI: -4.62, 38.34] per 20 ppb 3-day
31 avg03).
32 A study in Israel examined the association between pollutant concentrations and
33 fibrinogen among 3659 apparently healthy individuals (Steinvil et al.. 2008). In single
34 pollutant models, O3 was associated with an increase in fibrinogen at a 4-day lag among
35 men and a same-day O3 concentration among women but results for other lags (0 through
36 7 days) were mixed (some positive, some negative; none statistically significant).
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Inflammatory markers
1 Air pollution and inflammatory markers (C-reactive protein [CRP], white blood cell
2 [WBC] count, albumin, and Interleukin-6 [IL-6]) were also examined in several studies.
3 The ARIC study cohort, which included men and women aged 45-64 years old at the start
4 of the study, was utilized to assess the association between O3 concentrations and makers
5 of inflammation (Liao et al., 2005). No association was observed between O3
6 concentrations and albumin or WBC count.
7 Thompson et al. (2010) assessed ambient air pollution exposures and IL-6. This
8 retrospective repeated measures analysis was conducted among 45 adults (18-40 years of
9 age) in Toronto, Canada between the years of 1999 and 2006. Single pollutant models
10 were used to analyze the repeated-measures data using moving averages up to 7 days. A
11 positive association was observed between IL-6 and O3 with the strongest effects
12 observed for the 4-day moving average of O3 (quantitative results not provided). No
13 association was seen for shorter averaging times (<1 day). When examined by season
14 using 2-day moving averages, the association between O3 and IL-6 was positive during
15 only the spring and summer.
16 In Rotterdam, the Netherlands, a repeated measures study of healthy individuals living or
17 working in the city center reported no association between O3 concentration and CRP
18 (Rudez et al.. 2009). Each of the 40 participants provided between 11 and 13 blood
19 samples throughout a 1-year period (498 samples on 197 days). No consistent evidence of
20 an association was observed between O3 and CRP (percent change: -0.48 [95% CI: -
21 14.05, 13.10] per 20 ppb lag 1-day O3). Additionally, no association was observed with 2
22 or 3 day lags.
23 The relationship between pollutant concentrations and one-time measures of
24 inflammatory biomarkers was assessed in sex-stratified analyses among 3659 apparently
25 healthy individuals in Tel Aviv, Israel (Steinvil et al.. 2008). No evidence of an
26 association was observed between O3 and CRP or WBC for men and women.
27 A panel study of healthy individuals (n=76) was conducted in Taiwan to assess the
28 relationship between air pollutants and inflammation (Chuang et al.. 2007a). Health
29 endpoints were measured three times from April to June in 2004 or 2005. Ozone effects
30 were assessed in statistical models using the average of the 24 hours (1 day), 48 hours
31 (2 days), and 72 hours (3 days) before the hour of each blood sampling. Increases in CRP
32 were associated with increases in O3 concentrations in single-pollutant models (percent
33 change in CRP: 244.38 [95% CI: 4.54, 585.15] per 20 ppb 3-day avg O3). The association
34 was also observed using a 2-day averaging time, but no association was present with a 1-
3 5 day averaging time.
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Oxidative stress markers
1 A few studies have reported on the relationships between O3 concentration and oxidative
2 stress markers. The association between O3 exposure and markers of lipid peroxidation
3 and antioxidant capacity was examined among 120 nonsmoking healthy college students,
4 aged 18-22 years, from the University of California, Berkeley (February-June 2002)
5 (Chen et al.. 2007). By design, students were chosen that had experienced different
6 geographic concentrations of O3 over their lifetimes and during recent summer vacation
7 in either greater Los Angeles (LA) or the San Francisco Bay Area (SF). Long-term
8 (based on lifetime residential history) and shorter-term (based on the moving averages of
9 8-h max concentrations 1-30 days prior to the day of blood collection) O3 exposures were
10 estimated (lifetime exposure results presented in the chronic exposure section). A marker
11 of lipid peroxidation, 8-isoprostane (8-iso-PGF), was assessed. This marker is formed
12 continuously under normal physiological conditions but has been found at elevated
13 concentrations in response to environmental exposures. A marker of overall antioxidant
14 capacity, ferric reducing ability of plasma (FRAP), was also measured. Substantial
15 overlap in the more recent O3 exposure estimates (8-h moving averages) was observed
16 between the two geographic areas sampled. Levels of 8-iso-PGF were associated with
17 2-week ((3 = 0.035 [pg/mL]/8-h ppb O3, p = 0.007) and 1-month ((3 = 0.031 [pg/mL]/8-
18 h ppb O3, p = 0.006) estimated O3 exposure levels. No evidence of association was
19 observed between O3 and FRAP. A chamber study performed among a subset of study
20 participants supported the primary study results. The concentrations of 8-iso-PGF
21 increased immediately after the 4-h controlled O3 exposure ended (p = 0.10). However,
22 levels returned to near baseline by 18 hours without further exposure. The authors note
23 that O3 was highly correlated with PMi0-2.5 and NO2 in this study population; however,
24 inclusion of these pollutants in the O3 models did not substantially change the magnitude
25 of the associations with O3.
26 Using blood samples collected between April and June of 2004 or 2005 in Taiwan, the
27 association between O3 concentrations and a marker of oxidative stress was studied
28 among healthy individuals (n=76) (Chuang et al.. 2007a). Increases in 8-hydroxy-2'-
29 deoxyguanosine (8-OHdG) were associated with increases in O3 concentrations in single-
30 pollutant models (percent change in 8-OHdG: 2.46 [95% CI: 1.01, 3.92] per 20 ppb 1-day
31 avg O3). The association did not persist with 2- or 3-day averaging times.
Markers of overall cardiovascular health
32 Multiple studies used markers that assess overall cardiovascular well-being. Wellenius et
33 al. (2007) examined B-type natriuretic peptide (BNP), a marker of heart failure, in a
34 repeated-measures study conducted in Boston among 28 patients with congestive heart
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1 failure and impaired systolic function. The authors found no evidence of an association
2 between BNP and short-term O3 exposures at lags 0-3 days (quantitative results not
3 provided). BNP was chosen because it is directly associated with cardiac hemodynamics
4 and symptom severity among those with heart failure and is, therefore, considered a
5 marker of functional status. However, the authors conclude that the use of BNP may not
6 be useful in studies of the health effects of ambient air pollutants due to the large amount
7 of within-person variability in BNP levels observed in this population.
8 The relationship between air pollution and oxygen saturation and pulse rate, markers of
9 physiological well-being, was examined in a 2-month panel study among 31 congestive
10 heart failure patients (aged 50-85 years) in Montreal, Canada from July 2002 to October
11 2003 (Goldberg et al.. 2008). All participants had limited physical functioning
12 (New York Heart Association Classification > II) and an ejection fraction (the fraction of
13 blood pumped out of the heart per beat) less than or equal to 35% (normal is above 55%).
14 Daily mean O3 concentrations were calculated based on hourly measures at 10 monitoring
15 stations. There was an inverse association between O3 (lag-0) and oxygen saturation
16 when adjustment was made for temporal trends. In the models incorporating personal
17 covariates and weather factors, the association remained but was not statistically
18 significant. The associations of O3 with a lag of 1 day or a 3-day mean were not
19 statistically significant. No evidence of association was observed between O3 exposure
20 and pulse rate.
21 Total homocysteine (tHcy) is an independent risk factor for vascular disease and
22 measurement of this marker after oral methionine load is used to identify individuals with
23 mild impairment of homocysteine metabolism. The effects of air pollution on fasting and
24 postmethionine-load tHcy levels were assessed among 1,213 apparently healthy
25 individuals from Lombardia, Italy from January 1995 to September 2005 (Baccarelli et
26 al.. 2007). An increase in the 24-h O3 concentrations was associated with an increase in
27 fasting tHcy (percent change 6.25 [95% CI: 0.84, 11.91] per 20 ppb O3) but no
28 association was observed with postmethionine-load tHcy (percent change 3.36 [95% CI: -
29 1.30, 8.39] per 20 ppb O3). In addition, no evidence of association was observed between
30 7-day O3 concentrations and tHcy (percent change for fasting tHcy 4.16 [95% CI: -1.76,
31 10.42] and percent change for postmethionine-load tHcy -0.65 [95% CI: -5.66, 4.71] per
32 20 ppb O3). No evidence of effect modification by smoking was observed.
Blood lipids and glucose metabolism markers
33 Chuang et al. (2010) conducted a population-based cross-sectional analysis of data
34 collected on 7,778 participants during the Taiwanese Survey on Prevalence of
35 Hyperglycemia, Hyperlipidemia, and Hypertension in 2001. Apolipoprotein B (ApoB),
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1 the primary apolipoprotein among low-density lipoproteins, was associated with 3-day
2 avg O3 at the p < 0.10 level. The 5-day mean O3 concentration was associated with an
3 increase in triglycerides at p < 0.10. In addition, the 1-, 3-, and 5-day mean O3
4 concentrations were associated with increased HbAlc levels (a marker used to monitor
5 the degree of control of glucose metabolism) at the p < 0.05 level. The 5-day mean O3
6 was associated with increased fasting glucose levels (p < 0.10). No association was
7 observed between O3 concentration and ApoAl. Copollutant models were not assessed.
6.3.2.5 Myocardial Infarction (Ml)
8 The 2006 O3 AQCD did not report consistent results indicating an association between
9 short-term O3 exposure and MI. One study reported a positive association between
10 current day O3 concentration and acute MI, especially among the oldest age group (55- to
11 64-year olds) (Ruidavets et al.. 2005b). No association was observed in a case-crossover
12 study of O3 during the hours surrounding the event and MI (Peters etal.. 2001). Since the
13 2006 O3 AQCD, a few new epidemiologic studies have examined the association between
14 O3 exposure and MI (Henrotin et al., 2010; Rich et al.. 2010). as well as one study
15 published on arterial stiffness CWuet al.. 2010) and one study published on ST-segment
16 depression (Delfino et al., 2011).
17 One of the studies conducted in the U.S. examined hospital admissions for first MI and
18 reported no association with O3 concentrations (Rich et al.. 2010). More details on this
19 study are reported in the section on hospital admissions. Another study, performed in
20 Dijon, France, examined the association between O3 concentration and incident and
21 recurrent MI (Henrotin et al.. 2010). The mean 8-h O3 concentration was 19.1 ppb (SD
22 12.2 ppb). Odds ratios for the association between cumulative O3 concentrations and
23 recurrent Mis were elevated but none of the results were statistically significant (OR:
24 1.71 [95% CI: 0.91, 3.20] per 20 ppb for cumulative 1-3 day O3 exposure). No
25 association was observed for incident Mis. In analyses stratified by vascular risk factors,
26 positive associations were observed between 1-day lagged O3 concentrations and Mis
27 (incident and recurrent combined) among those who reported having
28 hypercholesterolaemia (OR: 1.52 [95% CI: 1.08, 2.15] per 20 ppb O3) and a slight inverse
29 association was observed among those who reported not having hypercholesterolaemia
30 (OR: 0.69 [95% CI: 0.50, 0.94] per 20 ppb O3). In other stratified analyses combining
31 different vascular factors, only those containing individuals with hypercholesterolaemia
32 demonstrated a positive association; none were inverse associations.
33 Wu et al. (2010) examined mail carriers aged 25-46 years and measured exposure to O3
34 through personal monitors [mean O3 24.9 (SD 14.0) ppb]. Ozone exposure was positively
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1
2
3
4
5
6
7
8
9
10
associated with arterial stiffness (percent change 11.24% [95% CI: 3.67, 19.62] per 40
ppb O3) and was robust to adjustment for ultrafme PM.
A study performed in the Los Angeles basin reported on the association between O3
exposure and ST-segment depression, a measure representing cardiac ischemia (Delfino
et al.. 2011). Study participants were nonsmokers, at least 65 years old, had a history of
coronary artery disease, and were living in a retirement community. Study periods
included five consecutive days in both July to mid-October and mid-October to February.
Mean 24-h O3 concentrations were 27.1 ppb (SD 11.5 ppb). No association was observed
between O3 concentrations and ST-segment depression of at least 1.0 mm during any of
the exposure periods (i.e., 1-h, 8-h, 1-day, 2-day avg, 3-day avg, 4-day avg).
11
12
13
14
6.3.2.6 Blood Pressure
In the 2006 O3 AQCD, no epidemiologic studies examined O3-related effects on blood
pressure (BP). Recent studies have been conducted to evaluate this relationship and
overall the findings are inconsistent. The O3 concentrations for these studies are listed in
Table 6-32.
Table 6-32 Characterization of ozone concentrations (in ppb) from studies of
blood pressure
Reference
Choi etal. (2007)
Delfino et al. (201 Ob)
Zanobetti et al. (2004)
Chuang etal. (2010)
Location
Incheon, South Korea
Los Angeles,
California
Boston,
Massachusetts
Taiwan
Averaging Time
8-h
(warm season)
8-h
(cold season)
24-h
1-h
5-days
Mean Concentration
(Standard Deviation)
26.6(11.8)
17.5(7.3)
27.1 (11.5)
20
24
26.83 (9.7)
Upper Range of
Concentration
75th: 34.8
Max: 62.4
75th: 22.9
Max: 33.9
Max: 60.7
Max: 62.1
15
16
17
18
19
20
21
Zanobetti et al. (2004) examined the relationship between air pollutants and BP from
May 1999 to January 2001 for 631 repeat visits among 62 Boston residents with CVD. In
single-pollutant models, higher resting diastolic blood pressure (DBP) was associated
with the 5-day (0-4 days) averages of O3 (RR: 1.03 [95% CI: 1.00, 1.05] per 20 ppb
increase in 24-h O3 concentrations). However, this effect was no longer apparent when
PM2 5 was included in the model (data were not presented) (Zanobetti et al.. 2004).
Delfino et al. (201 Ob) examined 64 subjects 65 years and older with coronary artery
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September 2011
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1 disease, no tobacco smoke exposure, and living in retirement communities in the
2 Los Angeles air basin with hourly (up to 14-h/day) ambulatory BP monitoring for 5 days
3 during a warm period (July-mid-October) and 5 days during a cool period (mid-October-
4 February). Investigators assessed lags of 1, 4, and 8 hours, 1 day, and up to 9 days before
5 each BP measure; no evidence of association was observed for O3 exposures (change in
6 BP associated with a 20 ppb change in 24-h O3 was 0.67 [95% CI: -1.16, 2.51 for systolic
7 BP [SBP] and -0.25 [95% CI: -1.25, 0.75] for DBP) (Delfino et al.. 2010b). Choi et al.
8 (2007) conducted a cross-sectional study to investigate the relationship between air
9 pollutants and BP among 10,459 participants of the Inha University Hospital health
10 examination from 2001 to 2003. These individuals had no medical history of
11 cardiovascular disease or hypertension. O3 exposure was associated with an increase in
12 SBP for 1-day lag in the warm season and similar effect estimates were observed during
13 the cold season but were not statistically significant (quantitative results not provided).
14 Associations between O3 and DBP were present in the cold season but not the warm
15 season (quantitative results not provided). The interaction term between O3 and season
16 was statistically significant. Chuang et al. (2010) conducted a similar type of study
17 among 7,578 participants of the Taiwanese Survey on Prevalence of Hyperglycemia,
18 Hyperlipidemia, and Hypertension in 2001. Investigators examined 1-, 3-, and 5-day avg
19 O3 concentrations. An increase in DBP was associated with the 3-day mean O3
20 concentration (change in BP for a 20 ppb increase in O3 was 0.61 [95% CI: 0.07, 1.14])
21 (Chuang et al.. 2010). Associations were not observed for other days or with SBP.
6.3.2.7 Hospital Admissions and Emergency Department Visits
22 Upon evaluating the collective evidence for O3-related cardiovascular hospital admissions
23 (HAs) and emergency department (ED) visits, the 2006 O3 AQCD concluded that "a few
24 studies observed positive O3 associations, largely in the warm season. Overall, however,
25 the currently available evidence is inconclusive regarding any association between
26 ambient O3 exposure on cardiovascular hospitalizations" (U.S. EPA. 2006b). Table 6-33
27 below provides information on the O3 concentrations reported in each of the recent HA
28 and ED visit studies evaluated.
29 Multiple recent studies of O3 exposure and cardiovascular HAs and ED visits have been
30 conducted in the U.S. and Canada. Peel et al. (2007) used a case-crossover framework
31 (using a time-stratified approach matching on day of the week in the calendar month of
32 the event) to assess the relationship between air pollutants and cardiovascular disease ED
33 visits among those with and without secondary comorbid conditions (hypertension,
34 diabetes, chronic obstructive pulmonary disease [COPD], congestive heart failure [CHF],
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Table 6-33 Characterization of ozone concentrations (in ppb) from studies of
HAs and ED visits
Study
Azevedoetal. (2011)
Ballesteretal. (2006)
Bell et al. (2008)
Buadong et al. (2009)
Cakmaketal. (2006a)
Chan etal. (2006)
Halonenetal. (2009)
Hosseinpoor et al. (2005)
Lanki etal. (2006)
Larrieuetal. (2007)
Lee et al. (2003b)
Lee et al. (2007)
Middleton et al. (2008)
Peel et al. (2007)
Rich etal. (2010)
Stieb et al. (2009)
Symons et al. (2006)
Tolbert etal. (2007)
Villeneuveetal. (2006a)
Von Klot etal. (2005)
Wellenius et al. (2005)
Yang (2008)
Zanobetti and Schwartz (2006)
Location
Portugal
Multicity, Spain
Taipei, Taiwan
Bangkok, Thailand
Multicity, Canada
Taipei, Taiwan
Helsinki, Finland
Tehran, Iran
Multicity, Europe
Multicity France
Seoul, Korea
Kaohsiung, Taiwan
Nicosia, Cyprus
Atlanta, GA
New Jersey
Multicity, Canada
Baltimore, MD
Atlanta, GA
Edmonton, Canada
Multicity, Europe
Allegheny County, PA
Taipei, Taiwan
Boston, MA
Averaging
Time
1-h
8-h
warm season
24-h
1-h
1-h max
1-h max
8-h max
warm season
8-h max
8-h max
warm season
8-h max
warm season
1-h max
24-h
8-h max
8-h max
warm season
24-h
24-h
8-h
warm season
8-h max
warm season
24-h
24-h
warm season
24-h
cold season
8h max
warm season
24-h
24-h
24-h
Mean Concentration
(Standard Deviation)
NR
24.2 - 44.3
21.4
14.4(3.2)
17.4
50.9 (26.4)
35.7*
4.9 (4.8)
31.7-57.2*
34.2-53.1
36.0(18.6)
26.5
28.7 - 54.9
55.6 (23.8)
NR
18.4
31.0(20.0)
53.0
17(9.1)
21.8(8)
12.2(7.4)
16.4-28.0
24.3(12.2)
21.0
22.4*
Upper Range of
Concentration
Max: 53.4
Max: 41 .9
Max: 150.3
75th: 42.1
Max: 79.6
75th: 7.2
Max: 99.0
75th: 44.9
75th: 35.5
Max: 83.0
Max: 120.0
75th: 67.0
Max: 147.5
75th: 23.5
75th: 27.0
75th: 17.0
75th: 32.0
75th: 26.3
Max: 62.8
75th: 31 .0
'Median presented (information on mean not given). NR: Not reported
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1 and dysrhythmia). Data on over 4 million ED visits from 31 hospitals were collected
2 from January 1993 to August 2000. Ozone was monitored from March to October. This
3 study was a re-analysis of a time series study conducted to assess the main effects of air
4 pollutants on cardiovascular ED visits in Atlanta (Tolbert et al.. 2007; Metzger et al..
5 2004). In the initial study, no evidence of associations was observed between O3 and all
6 CVD visits or visits for CVD subgroups, such as dysrhythmia, CHF, ischemic heart
7 disease (HD), and peripheral vascular and cerebrovascular disease. The relative risk for
8 all CVD visits was 1.01 (95% CI: 0.99, 1.02) for a 20 ppb increase in the 3-day moving
9 avg (lags 0-2 days) of 8-h O3 (Metzger et al., 2004). Similar to the initial investigation
10 using a time-series analysis, no evidence of an association was observed for the O3 3-day
11 moving average and CVD visits among the entire population using the case-crossover
12 design (Peel et al.. 2007). However, the relationship between O3 and peripheral and
13 cerebrovascular disease visits was stronger among patients with comorbid COPD (OR:
14 1.19 [95% CI: 1.03-1.36] per 20 ppb, lag 0-2 days) as compared to patients without
15 COPD (OR: 1.01 [95% CI: 0.97-1.04] per 20 ppb, lag 0-2 days). The same research
16 group expanded upon the number of Atlanta hospitals providing ED visit data (41
17 hospitals) as well as the length of the study period (1993-2004) (Tolbert et al.. 2007).
18 Again, models assessing the health effects of O3 utilized data collected from March
19 through October. Similar to the results presented by Metzger et al. (2004) and Peel et al.
20 (2007) among the entire study population, no evidence of associations was observed for
21 O3 and CVD visits (Tolbert et al.. 2007).
22 A study of HAs for MI was performed using a statewide registry from New Jersey
23 between January 2004 and December 2006 (Rich et al.. 2010). Using a case-crossover
24 design, the association between the previous 24 hr O3 concentration and transmural
25 infarction (n=l,003) was examined. No association was observed (OR: 0.94 [95% CI:
26 0.79, 1.13] per 20 ppb) and this did not change with the inclusion of PM25 in the model.
27 Cakmak et al, (2006b) investigated the relationship between gaseous air pollutants and
28 cardiac hospitalizations in 10 large Canadian cities using a time-series approach. A total
29 of 316,234 hospital discharge records for primary diagnosis of congestive heart failure,
30 ischemic heart disease, or dysrhythmia were obtained from April 1993 through March
31 2000. Correlations between pollutants varied substantially across cities, which could
32 partially explain discrepancies in effect estimates observed across the cities. In addition,
33 pollutant lags differed across cities; the average lag for O3 was 2.9 days. The pooled
34 effect estimate for a 20 ppb increase in the daily 1-h max O3 concentration and the
35 percent change in hospitalizations among all 10 cities was 2.3 (95% CI: 0.11, 4.50) in an
36 all-year analysis. The authors reported no evidence of effect modification by gender,
37 neighborhood-level education, or neighborhood-level income. A similar multicity time-
38 series study was conducted using nearly 400,000 ED visits to 14 hospitals in seven
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1 Canadian cities from 1992 to 2003 (Stieb et al., 2009). Primary analyses considered daily
2 O3 single day lags of 0-2 days; in addition, sub-daily lags of 3-h avg concentrations up to
3 12 hours before presentation to the ED were considered. Seasonal variation was assessed
4 by stratifying analyses by warm and cold seasons. No evidence of effect of O3 on CVD
5 ED visits was observed. One negative, statistically significant association was reported
6 between a 1-day lag of O3 and visits for angina/myocardial infarction. Ozone was
7 negatively correlated with many of the other pollutants, particularly during the cold
8 season.
9 The effect of air pollution on daily ED visits for ischemic stroke (n=10,881 visits) in
10 Edmonton, Canada was assessed from April 1992 through March 2002 (Szyszkowicz.
11 2008). A 26.4% (95% CI: 3.16-54.5) increase in stroke ED visits was associated with a
12 20 ppb increase in O3 at lag 1 among men aged 20-64 years in the warm season. No
13 associations were present among women or among men age 65 and older. In addition, no
14 associations were observed for the cold season or for other lags (lag 0 or lag 2). A similar
15 investigation over the same time period in Edmonton, Canada, assessed the relationship
16 between air pollutants and ED visits for stroke (ischemic stroke, hemorrhagic stroke, and
17 transient ischemic attack) among those 65 years of age and older using a case-crossover
18 framework (Villeneuve et al.. 2006a). Two-pollutant models were assessed. No evidence
19 of association was reported for O3 and stroke hospitalization (Villeneuve et al.. 2006a).
20 Additional studies reported no evidence of an association between O3 concentrations and
21 ED visits, hospitalizations, or symptoms leading to hospitalization (Symons et al.. 2006;
22 Zanobetti and Schwartz. 2006; Wellenius et al., 2005). Symons et al. (2006) used a case-
23 crossover framework to assess the relationship between air pollutants and the onset of
24 symptoms (dyspnea) severe enough to lead to hospitalization (through the ED) for
25 congestive heart failure. The study was conducted from April to December of 2002 in
26 Baltimore, Maryland. Exposures were assigned using 3 index times: 8-h and 24-h periods
27 prior to symptom onset and date of hospital admission. No evidence of association was
28 reported for O3 concentrations. Although seasonal variation was not assessed, the time
29 frame for the study did not involve an entire year (April to December). Wellenius et al.
30 (2005) investigated the association between air pollutants and congestive heart failure
31 hospitalization among Medicare beneficiaries in Pittsburgh, Pennsylvania from 1987 to
32 1999 utilizing a case-crossover framework. A total of 55,019 admissions from the
33 emergency room with a primary discharge diagnosis of CHF were collected. No evidence
34 of an association was reported for O3 and CHF hospitalization (Wellenius et al., 2005).
35 Finally, Zanobetti and Schwartz (2006) assessed the relationship between air pollutants
36 and hospital admissions through the ED for myocardial infarction and pneumonia among
37 patients aged 65 and older residing in the greater Boston area (1995-1999) using a case-
38 crossover framework with control days in the same month matched on temperature.
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1 Pollution exposures were assigned for the same day and for the mean of the exposure the
2 day of and the day before the admission. Ozone was not associated with MI admissions in
3 all-year and seasonal analyses.
4 Several recent studies have examined the relationship between air pollution and CVD
5 hospital admissions and/or emergency department visits in Asia. Of note, some areas of
6 Asia have a more tropical climate than the U.S. and do not experience similar seasonal
7 changes. In Taiwan, fairly consistent positive associations have been reported for O3 and
8 congestive heart failure hospital admissions (for single- and copollutant models) in Taipei
9 on warm days (Yang. 2008) and in Kaohsiung (Lee et al.. 2007): cerebrovascular disease
10 ED visits (for lag 0 single- and two-pollutant models but not other lags or 3-pollutant
11 models) in Taipei (Chan et al.. 2006); and arrhythmia ED visits in Taipei among those
12 without comorbid conditions (Chiu et al.. 2009: Lee et al.. 2008a) and in Taipei on warm
13 days among those with and without comorbid conditions (Lee et al.. 2008a: Jansson et al..
14 2001). However, one study in Taiwan did not shown an association. Bell et al. (2008)
15 reported no evidence of an O3 association with hospital admissions for ischemic heart
16 disease or cerebrovascular disease. Three studies based in Asia but outside Taiwan were
17 performed. First, a Hong Kong-based investigation (Wong et al.. 2009) reported no
18 consistent evidence of a modifying effect of influenza on the relationship between O3 and
19 CVD admissions. Second, among elderly populations in Thailand, O3 was associated with
20 CVD visits, but this association was not detected among younger age groups (15-64)
21 (Buadong et al.. 2009). Third, a study performed in Seoul, Korea reported a positive
22 association between O3 levels and HAs for ischemic heart disease; the association was
23 slightly greater among those over 64 years of age (Lee etal.. 2003b).
24 Positive effects of O3 on CVD hospital admissions and/or ED visits have been reported in
25 other areas of the world as well (Azevedo et al.. 2011: Linares and Diaz. 2010: Middleton
26 et al.. 2008: Turner et al.. 2007: Yallop et al.. 2007: Ballester et al.. 2006: De Pablo et al..
27 2006: Von Klot et al.. 2005). although not consistently as some studies reported no
28 association (Oudin et al.. 2010: Halonen et al.. 2009: Larrieu et al.. 2007: Barnett et al..
29 2006: Hinwood et al.. 2006: Lanki et al.. 2006: Hosseinpoor et al.. 2005: Simpson et al..
30 2005).
31 A couple of studies (U.S. and Australia) have examined cardiac arrests where emergency
32 services attempted treatment/resuscitation. No evidence of an association between O3 and
33 out-of-hospital cardiac arrest was observed (Dennekamp et al.. 2010: Silverman et al..
34 2010).
3 5 An increasing number of air pollution studies have investigated the relationship between
36 O3 concentrations and CVD hospital admissions and/or ED visits. As summarized in the
37 2006 O3 AQCD, some, especially those reporting results stratified by season (or
Draft - Do Not Cite or Quote 6-165 September 2011
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1
2
3
4
5
6
7
temperature) or comorbid conditions have reported positive associations. However, even
studies performing these stratified analyses are not consistent and the overall evidence
remains inconclusive regarding the effects of O3 on CVD HAs and ED visits. These HA
and ED visit studies are summarized in Figures 6-22 through 6-26, which depict the
associations for studies in which numerical associations were presented for an overall
study population. Tables 6-34 through 6-38 provide the numerical results displayed in the
figures.
Reference
Buadongetal. (2009)
Middleton et al. (2008)
Fungetal. (2005)
Ballesteretal. (2001)
Petroeschevskyetal. (2001)
Linn et al. (2000)
Atkinson etal. (1999)
Wongetal. (1999a)
Wong etal. (1999b)
Prescottetal. (1998)
Polonieckietal. (1997)
Halonen et al. (2009)
Larrieu et al. (2007)
Peel etal. (2007)
Ballesteretal. (2006)
Chang et al. (2005)
Yang et al. (2004)
Wongetal. (1999b)
Chang et al. (2005)
Yang et al. (2004)
Wongetal. (1999a)
Wongetal. (1999b)
Cakmak et al. (2006)
Ballesteretal. (2001)
Morgan etal. (1998)
Larrieu et al. (2007)
Ballesteretal. (2006)
von Klot et al. (2005)
Bell et al. (2008)
Chan et al. (2006)
Ballesteretal. (2001)
Wongetal. (1999a)
Wongetal. (1999b)
Polonieckietal. (1997)
Peel etal. (2007)
Wongetal. (1999b)
Wongetal. (1999b)
Location
Bangkok, Thailand
Nicosia, Cyprus
Brisbane, Australia •
Los Angeles, CA •
London, England
Hong Kong
Hong Kong
London, England 0
8 French cities -t
Atlanta, GA -4
14 Spanish cities
Taipei, Taiwan
Kaohsiung, Taiwan
Taipei, Taiwan
Kaohsiung, Taiwan
Hong Kong
Hong Kong
10 Canadian cities
Sydney, Australia
8 French cities —
14 Spanish cities
5 European cities
Taipei, Taiwan
Hong Kong 9
London, England 0
Atlanta, GA —
Hong Kong
•
»-
K
— o
-•-
•-
•
-•-
-o
Cardiovascular
disease
Cardiac disease
Cerebrovascular
disease
0.7
0.8 0.9
1 1.1
1.2
1.3 1.4 1.5
Note: Increase in O3 standardized to 20 ppb for 24-h avg period, 30 ppb for 8-h avg period, and 40 ppb for 1-h avg period. Ozone
concentrations in ppb. Seasons depicted by colors - black: all year; red: warm season; light blue: cold season. Age groups of study
populations were not specified or were adults with the exception of Fung et al. (2005). Wong et al. (1999b). and Prescott et al.
(1998). which included only individuals aged 65+.
Figure 6-22 Odds ratio (95% Cl) per increment ppb increase in ozone for over
all cardiovascular ED visits or HAs.
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September 2011
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Table 6-34 Odds ratio (95% Cl) per increment ppb increase in ozone for overall
cardiovascular ED visits or HAs in studies presented in Figure 6-22
Study
Atkinson et al. (2006a)
Ballesteretal. (2006)
Ballesteretal. (2006)
Bell et al. (2008)
Buadong et al. (2009)
Cakmaketal. (2006a)
Chan etal. (2006)
Chang et al. (2005)
Funa etal. (2006a)
Halonen et al. (2009)
Larrieu et al. (2007)
Linn et al. (2006a)
Middleton etal. (2008)
Morgan et al. (2008)
Peel et al. (2007)
Petroeschevskv etal. (2001)
Poloniecki et al. (2006a)
Prescott etal. (1998)
Von Klot etal. (2005)
Wong et al. (1999b)
Wong et al. (1999a)
Yang et al. (2005)
Location
London, England
Multicity, Spain
Valencia, Spain
Taipei, Taiwan
Bangkok, Thailand
Multicity, Canada
Taipei, Taiwan
Taipei, Taiwan
Wndsor, Canada
Helsinki, Finland
Multicity France
Los Angeles, California
Nicosia, Cyprus
Sydney, Australia
Atlanta, GA
Brisbane, Australia
London, England
Edinburgh, Scotland
Multicity, Europe
Hong Kong
Hong Kong
Kaohsiung, Taiwan
Outcome
Cardiovascular disease
Cardiovascular disease
Cardiac disease
Cardiovascular disease
Cardiac disease
Cerebrovascular disease
Cerebrovascular disease
Cardiovascular disease
Cardiac disease
Cerebrovascular disease
Cardiovascular disease
Cardiovascular disease
Cardiovascular disease
Cardiac disease
Cardiovascular disease
Cardiovascular disease
Cardiac disease
Cardiovascular disease
Cerebrovascular disease
Cardiovascular disease
Cardiovascular disease
Cerebrovascular disease
Cardiovascular disease
Cardiac disease
Cardiovascular disease
Cerebrovascular disease
Cardiovascular disease
Cerebrovascular disease
Cardiovascular disease
Averaging Time
8-h
8-h warm season
8-h warm season
8-h
8-h
8-h
24-h
1-h
1-h max
1-h max
24-h warm season
24-h cold season
1-h
8-h max warm season
8-h max warm season
24-h
8-h max
1-h max
8-h warm season
8-h warm season
8-h
8-h
8-h
24-h
8-h max warm season
24-h
24-h cold season
24-h
24-h
24-h warm season
24-h cold season
24-h
24-h warm season
24-h cold season
24-h warm season
24-h cold season
Standardized Estimate (95% Cl)
1.03(1.00,1.05)
1.04(1.02,1.06)
1.04(1.01,1.07)
0.94(0.84,1.06)
0.88(0.75,1.03)
0.86(0.72,1.04)
0.94(0.87,1.02)
1.01 (1.00, 1.02)
1.02(1.00,1.04)
1.02(1.01,1.03)
1.42(1.33,1.50)
1.15(1.04,1.27)
1.02(0.92,1.13)
1.05(0.96,1.14)
1.01 (0.98, 1.04)
0.99(0.98,1.00)
1.09(1.00,1.18)
1.02(0.99,1.05)
1.00(0.98,1.02)
1.02(0.98,1.05)
0.96(0.92,1.01)
0.97(0.93,1.01)
0.98(0.95,1.02)
0.89(0.78,1.00)
1.11 (1.00, 1.22)
1.08(1.03,1.13)
1.15(1.04,1.26)
0.95(0.90,1.01)
1.02(1.03,1.06)
1.01 (0.96, 1.06)
1.06(1.02,1.11)
0.99(0.95,1.04)
0.98(0.90,1.08)
1.02(0.96,1.10)
1.33(1.26,1.40)
1.05(0.96,1.15)
Note: Increase in 03 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h averaging period.
Ozone concentrations in ppb. Age groups of study populations were not specified or were adults with the exception of Fung et al. (2006a), Wong et
al. (1999a), and Prescott et al. (1998). which included only individuals aged 65+.
Warm season defined as: March-October (Peel etal.. 2007). May-October (Ballesteretal.. 2005: Wong etal.. 1999a). May-September (Halonen
etal.. 2009). April-September (Larrieu etal.. 2007: Von Klot et al.. 2005). > 20°C (Chang etal.. 2005) and > 25°C (Yang etal.. 2004). Cold season
defined as: November-April (Wong etal.. 1999a). <20°C (Chang etal.. 2005) and <25°C (Yang etal.. 2004). December-March (Wong etal.. 1999b)
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September 2011
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Reference
Stiebetal. (2009)
Welleniusetal. (2005)
Wongetal. (1999a)
Wongetal. (1999b)
Poloniecki et al. (1997)
Yang (2008)
Peel et al. (2007)
Lee et al. (2007)
Symonset al. (2006)
Wongetal. (1999b)
Yang (2008)
Lee et al. (2007)
Wongetal. (1999b)
Location
7 Canadian cities
Allegheny county, PA
Hong Kong
Hong Kong —
London, England
Taipei, Taiwan
Atlanta, GA
Kaohsiung, Taiwan
Baltimore, MD —
Hong Kong
Taipei, Taiwan
Kaohsiung, Taiwan
Hong Kong
0.4
0.6
0.8
1
1.2
1.4
1.6
1.8
Note: Increase in O3 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h
averaging period. Ozone concentrations in ppb. Seasons depicted by colors: black: all year; red: warm season; light blue: cold
season. Outcomes were all congestive heart failure, with the exception of Symons et al. (2006). which examined onset of congestive
heart failure symptoms leading to a heart attack. Age groups of study populations were not specified or were adults with the
exception of Wellenius et al. (2005) and Wong et al. (1999a). which included only individuals aged 65+.
Figure 6-23 Odds Ratio (95% Cl) per increment ppb increase in ozone for
congestive heart failure ED visits or HAs.
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September 2011
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Table 6-35 Odds Ratio (95% Cl) per increment ppb increase in ozone for
congestive heart failure ED visits or HAs for studies in Figure 6-23
Study
Lee et al. (2007)
Peel et al. (2007)
Poloniecki
etal. (1997)
Stieb et al. (2009)
Symons et al. (2006)
Wellenius et al. (2005)
Wong etal. (1999a)
Yang (2008)
Wonaetal. (1999b)
Location
Kaohsiung, Taiwan
Atlanta, GA
London, England
Multicity, Canada
Baltimore, MD
Allegheny county, PA
Hong Kong
Taipei, Taiwan
Hong Kong
Outcome
congestive heart failure
congestive heart failure
congestive heart failure
congestive heart failure
congestive heart failure
onset of congestive heart failure
symptoms leading to heart attack
congestive heart failure
congestive heart failure
congestive heart failure
congestive heart failure
congestive heart failure
Averaging Time
24-h warm season
24-h cold season
8-h warm season
8-h
24-h
8-h warm season
24-h
24-h
24-h warm season
24-hcold season
24-h warm season
24-h cold season
24-h
Standardized
Estimate (95% Cl)
1.25(1.15,1.36)
1.24(1.09,1.41)
0.96(0.93,1.00)
0.99(0.95,1.03)
1.03(0.98,1.07)
0.83(0.49,1.41)
0.98(0.96,1.01)
1.11 (1.04,1.80)
1.09(0.96,1.23)
1.16(1.06,1.27)
1.39(1.27,1.51)
0.61 (0.52, 0.73)
1.25(1.11,1.41)
Note: Increase in 03 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h averaging
period. Ozone concentrations in ppb. Outcomes were all congestive heart failure, with the exception of Symons et al. (2006). which
examined onset of congestive heart failure symptoms leading to a heart attack. Age groups of study populations were not specified or were
adults with the exception of Wellenius et al. (2005) and Wong et al. (1999a). which included only individuals aged 65+.
Warm season defined as: March-October (Peel etal.. 2007). April-November (Svmons etal.. 2006). May-October (Wonaetal.. 1999a)
> 20°C (Yang. 2008). and >25°C (Lee etal.. 2007). Cold season defined as: November-April (Wong etal.. 1999a). <20°C (Yang. 2008). and
<25°C (Lee etal..2007).
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September 2011
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Reference
Buadong et al. (2009)
Bell et al. (2008)
Lee et al. (2003)
Atkinson etal. (1999)
Wongetal. (1999a)
Wong etal. (1999b)
Larrieu et al. (2007)
Peel et al. (2007)
Lee et al. (2003)
Wongetal. (1999b)
Wongetal. (1999b)
Location
Bangkok, Thailand
Taipei, Taiwan
Seoul, Korea
London, England
Hong Kong
Hong Kong
8 French cities
Atlanta, GA
Seoul, Korea
Hong Kong
Hong Kong
Halonen et al. (2009) Helsinki, Finland
Rich etal. (2010)
Buadong et al. (2009)
Stiebetal. (2009)
Zanobetti et al. (2006)
Poloniecki et al. (1997)
Lanki et al. (2006)
von Klot et al. (2005)
Hosseinpoor et al. (2005)
Poloniecki et al. (1997)
von Klot et al. (2005)
New Jersey
Bangkok, Thailand
7 Canadian cities
Boston, MA
London, England
5 European cities
5 European cities
Tehran, Iran
London, England
5 European cities
0.5
0.7
0.9
Ischemia heart disease
Coronary heart disease
Myocardial infarction
Angina pectoris
1.1
1.3
1.5
Note: Increase in O3 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h
averaging period. Ozone concentrations in ppb. Seasons depicted by colors: black: all year; red: warm season; light blue: cold
season. Age groups of study populations were not specified or were adults with the exception of Wong et al. (1999a) and Atkinson et
al. (2006a), which included only individuals aged 65+.
Figure 6-24 Odds Ratio (95% confidence interval) per increment ppb increase in
ozone for ischemic heart disease, coronary heart disease,
myocardial infarction, and angina pectoris ED visits or HAs.
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Table 6-36 Odds Ratio (95% Cl) per increment ppb increase in ozone for
ischemic heart disease, coronary heart disease, myocardial
infarction, and angina pectoris ED visits or HAs for studies
presented in Figure 6-24
Study
Atkinson et al. (1999)
Bell et al. (2008)
Buadong et al. (2009)
Halonenetal. (2009)
Hosseinpoor et al. (2005)
Lankietal. (2006)
Larrieuetal. (2007)
Lee et al. (2003b)
Peel et al. (2007)
Polonieckietal. (1997)
Rich et al. (Rich etal., 2010)
Stieb et al. (2009)
Von Klot etal. (2005)
Wong etal. (2009)
Wong etal. (2008)
Zanobetti and Schwartz (2006)
Location
London, England
Taipei, Taiwan
Bangkok, Thailand
Helsinki, Finland
Tehran, Iran
Multicity, Europe
Multicity France
Seoul, Korea
Atlanta, GA
London, England
New Jersey
Multicity, Canada
Multicity, Europe
Hong Kong
Hong Kong
Boston, MA
Outcome
Ischemic heart disease
Ischemic heart disease
Ischemic heart disease
Myocardial infarction
Coronary heart disease
Angina
Myocardial infarction
Ischemic heart disease
Ischemic heart disease
Ischemic heart disease
Ischemic heart disease
Myocardial infarction
Angina
Myocardial infarction
Myocardial infarction
Myocardial infarction
Angina
Ischemic heart disease
Ischemic heart disease
Myocardial infarction
Averaging Time
8-h
24-h
1-h
1-h
8-h max warm season
8-h max
8-h max warm season
8-h max warm season
1-h max
1-h max warm season
8-h warm season
8-h
8-h
24-h
2-h
8-h max warm season
8-h max warm season
24-h
24-h warm season
24-h cold season
24-h
24-h
Standardized
Estimate (95% Cl)
0.97(0.94,1.01)
1.01(0.91,1.12)
1.00(0.98,1.02)
0.97(0.94,1.01)
0.99 (0.79, 1 .25)
0.80 (0.70, 0.92)
0.96(0.92,1.01)
1 .02 (0.98, 1 .07)
1.07(1.02,1.13)
1.07(1.00,1.17)
1 .00 (0.97, 1 .03)
0.98(0.94,1.02)
0.98 (0.94, 1 .03)
0.94(0.79,1.13)
1 .00 (0.96, 1 .04)
1.00(0.83,1.21)
1.19(1.05,1.35)
1.01(0.94,1.06)
1.02(0.94,1.11)
1 .02 (0.95, 1 .09)
1 .03 (0.98, 1 .08)
0.98(0.92,1.03)
Note: Increase in 03 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h averaging period.
Ozone concentrations in ppb. Age groups of study populations were not specified or were adults with the exception of Wong et al. (1999a) and
Atkinson et al. (2006a). which included only individuals aged 65+.
Warm season defined as: March-October (Peel etal.. 2007). June-August (Lee etal.. 2003b). May-September (Halonenetal.. 2009). May-
October (Buadong et al.. 2009). and April-September (Larrieu etal.. 2007: Lanki et al.. 2006: Von Klot etal.. 2005). Cold season defined as:
November-April (Buadong etal.. 2009)
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Reference
Chan et al. (2006)
Halonen et al. (2009)
Larrieu et al. (2007)
Chan et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Chan et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Villeneuve et al. (2006)
Location
Taipei, Taiwan
Helsinki, Finland
8 French cities
Taipei, Taiwan
Edmonton, Canada
Edmonton, Canada
Edmonton, Canada
Taipei, Taiwan
Edmonton, Canada
Edmonton, Canada
Edmonton, Canada
Edmonton, Canada
Edmonton, Canada
Edmonton, Canada
All
Ischemia
Hemorrhagic
-O-
Transient
ischemic
-O
0.5
0.7
0.9
1.1
1.3
1.5
Note: Increase in O3 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h
averaging period. Ozone concentrations in ppb. Seasons depicted by colors: black: all year; red: warm season; light blue: cold
season. Age groups of study populations were not specified or were adults with the exception of Villeneuve et al. (2006a), which
included only individuals aged 65+, and Chan et al. (2006). which included only individuals aged 50+.
Figure 6-25 Odds Ratio (95% confidence interval) per increment ppb increase in
ozone for stroke ED visits or HAs.
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September 2011
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Table 6-37 Odds Ratio (95% Cl) per increment ppb increase in ozone for stroke
ED visits or HAs for studies presented in Figure 6-25
Study
Chan etal. (2006)
Halonenetal. (2009)
Larrieu etal. (2007)
Villeneuveetal. (2006a)
Location
Taipei, Taiwan
Helsinki, Finland
Multicity, France
Edmonton,
Canada
Outcome
All/non-specified stoke
Ischemia stroke
Hemorrhagic stroke
All/non-specified stoke
All/non-specified stoke
Ischemic stroke
Ischemic stroke
Ischemic stroke
Hemorrhagic stroke
Hemorrhagic stroke
Hemorrhagic stroke
Transient ischemic stroke
Transient ischemic stroke
Transient ischemic stroke
Averaging Time
1-hmax
1-h max
1-h max
8-h max warm season
8-h max warm season
24-h
24-h warm season
24-h cold season
24-h
24-h warm season
24-h cold season
24-h
24-h warm season
24-h cold season
Standardized Estimate
(95% Cl)
1.01(0.99,1.03)
1.03(0.99,1.07)
0.99(0.92,1.06)
1.08(0.83,1.41)
0.98 (0.93 , 1 .02)
1.00(0.88,1.13)
1.09(0.91,1.32)
0.98(0.80,1.18)
1 .02 (0.87, 1 .20)
1.12(0.88,1.43)
0.97 (0.76, 1 .22)
0.98(0.87,1.10)
0.85(0.70,1.01)
1.11 (0.93,1.32)
Note: Increase in 03 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h averaging period.
Ozone concentrations in ppb. Age groups of study populations were not specified or were adults with the exception ofVilleneuve etal. (2006a).
which included only individuals aged 65+, and Chan et al. (2006). which included only individuals aged 50+.
Warm season defined as: May-September (Halonenetal., 2009), and April-September (Larrieu etal., 2007: Villeneuve etal., 2006a). Cold
season defined as: October-March (Villeneuve et al.. 2006a).
Reference
Stieb et al. (2009)
Peel et al. (2007)
Location
7 Canadian cities
Atlanta, GA
Buadong et al. (2009) Bangkok, Thailand
Hong Kong
London, England
Wong etal. (1999b)
Polonieckiet al.
(1997)
Halonen et al. (2009) Helsinki, Finland
Wong etal. (1999b)
Wong etal. (1999b)
Hong Kong
Hong Kong
0.7
0.8
0.9
Dysrhythmia
Arrhythmia
1.1 1.2 1.3 1.4
Note: Increase in O3 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h
averaging period. Ozone concentrations in ppb. Seasons depicted by colors: black: all year; red: warm season; light blue: cold
season. Age groups of study populations were not specified or were adults with the exception of Wong et al. (1999a). which included
only individuals aged 65+.
Figure 6-26 Odds Ratio (95% confidence interval) per increment ppb* increase
in ozone for arrhythmia and dysrhythmia ED visits or HAs.
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Table 6-38 Odds Ratio (95% Cl) per increment ppb* increase in ozone for
arrhythmia and dysrhythmia ED visits or HAs for studies presented
in Figure 6-26
Study Location Outcome Averaging Time
Buadonq et al. (2009)
Halonenetal. (2009)
Peel et al. (2007)
Polonieckietal. (2009)
Stieb et al. (2009)
Wong etal. (2009)
Bangkok, Thailand
Helsinki, Finland
Atlanta, GA
London, England
Multicity, Canada
Hong Kong
Arrhythmia
Arrhythmia
Dysrhythmia
Arrhythmia
Dysrhythmia
Arrhythmia
1-h
8-h max warm season
8-h warm season
8-h
24-h
24-h
24-h warm season
24-h cold season
0.99 (0.95,
1 .04 (0.80,
1 .01 (0.98,
1 .02 (0.96,
1 .02 (0.95,
1 .06 (0.99,
1.10(0.96,
1.11 (1.01,
1.04)
1.35)
1.04)
1.07)
1.09)
1.12)
1.26)
1.23)
Note: Increase in 03 standardized to 20 ppb for 24-h averaging period, 30 ppb for 8-h averaging period, and 40 ppb for 1-h averaging period.
Ozone concentrations in ppb. Age groups of study populations were not specified or were adults with the exception of (Wong etal., 1999a), which
included only individuals aged 65+. Warm season defined as: March-October (Peel etal..2007). May-October (Wong etal.. 1999a) and May-
September (Halonenetal..2009). Cold season defined as: November-April (Wong etal.. 1999a).
6.3.2.8 Cardiovascular Mortality
1 As discussed within this section (Section 6.3), epidemiologic studies provide inconsistent
2 evidence of an association between short-term O3 exposure and cardiovascular effects.
3 However, toxicological studies have demonstrated O3-induced cardiovascular effects,
4 specifically enhanced atherosclerosis and ischemia, which could lead to death. The 2006
5 O3 AQCD
6 provided evidence, primarily from single-city studies, of consistent positive associations
7 between short-term O3 exposure and cardiovascular mortality. Recent multicity studies
8 conducted in the U.S., Canada, and Europe further confirm the association between short-
9 term O3 exposure and cardiovascular mortality.
10 As discussed in Section 6.2.7.2, the APHENA study (Katsouyanni et al.. 2009) also
11 examined associations between short-term O3 exposure and mortality and found
12 consistent positive associations for cardiovascular mortality in all-year analyses with
13 associations persisting in analyses restricted to the summer season. Additional multicity
14 studies from the U.S. (Zanobetti and Schwartz. 2008b). Europe (Samoli et al.. 2009).
15 Italy (Stafoggia et al., 2010). and Asia (Wong et al.. 2010) that conducted summer season
16 and/or all-year analyses provide additional support for an association between short-term
17 O3 exposure and cardiovascular mortality (Figure 6-37).
18 Of the studies evaluated, only the APHENA study (Katsouyanni et al.. 2009) and the
19 Italian multicity study (Stafoggia et al.. 2010) conducted an analysis of the potential for
20 copollutant confounding of the O3-cardiovascular mortality relationship. In the European
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1 dataset, when focusing on the natural spline model with 8 df/year (Section 6.2.7.2) and
2 lag 1 results in order to compare results across study locations (Section 6.6.2.1),
3 cardiovascular mortality risk estimates were robust to the inclusion of PM10 in
4 copollutant models in all-year analyses with more variability in the Canadian and U.S.
5 datasets (i.e., cardiovascular O3 mortality risk estimates were reduced or increased in
6 copollutant models). In summer season analyses, cardiovascular O3 mortality risk
7 estimates were robust in the European dataset and attenuated but remained positive in the
8 U.S. dataset. Similarly, in the Italian multicity study (Stafoggia et al.. 2010). which was
9 limited to the summer season, cardiovascular mortality risk estimates were robust to the
10 inclusion of PMi0 in copollutant models. Based on the APHENA and Italian multicity
11 results, O3 cardiovascular mortality risk estimates appear to be robust to inclusion of
12 PMio in copollutant models. However, in the U.S. and Canadian datasets there was
13 evidence that O3 cardiovascular mortality risk estimates are moderately to substantially
14 sensitive (e.g., increased or attenuated) to PMi0. The mostly every-6th-day sampling
15 schedule for PM10 in the Canadian and U.S. datasets greatly reduced their sample size
16 and limits the interpretation of these results.
6.3.2.9 Summary of Epidemiologic Studies
17 Overall, the available body of evidence examining the relationship between short-term
18 exposures to O3 and cardiovascular morbidity is inconsistent. Differences in exposure
19 metrics and windows of exposure, a wide variety of biomarkers considered, and a lack of
20 consistency among definitions used for specific cardiovascular disease endpoints (e.g.
21 arrhythmias, HRV) make comparisons across studies difficult. In addition, several
22 investigators reporting associations between O3 and cardiovascular morbidity postulate
23 that O3 may be acting as a proxy for sulfate; differences reported across multicity studies
24 and across studies conducted in specific cities/regions point to the importance of
25 considering multipollutant relationships that vary across geographic regions. Additionally
26 mortality studies indicate a consistent positive association between O3 and cardiovascular
27 mortality.
6.3.3 Toxicology
6.3.3.1 Summary of Findings from Previous Ozone AQCDs
28 In the previous O3 AQCDs (U.S. EPA. 2006b. 1996a) experimental animal studies have
29 reported relatively few cardiovascular system alterations after exposure to O3 and other
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1 photochemical oxidants. The limited amount of research directed at examining O3-
2 induced cardiovascular effects has primarily found alterations in heart rate (HR) and BP
3 after O3 exposure. A group of studies (Arito et al.. 1992; Arito etal.. 1990; Uchiyama and
4 Yokovama. 1989; Yokovama et al.. 1989; Uchivama et al.. 1986) report Q3 (0.1-1.0 ppm)
5 exposure in rats decreased core temperature (Tco), HR, and mean arterial pressure
6 (MAP). However, these cardiovascular responses to O3 could be attenuated by increased
7 ambient temperatures, exhibited adaptation, and were the result of the rodent
8 hypothermic response (Watkinson et al.. 2003; Watkinson et al.. 1993). This hypothermic
9 response could be an attempt to minimize the irritant effects of O3 inhalation, serving as a
10 physiological and behavioral defense mechanism (Twasaki etal.. 1998; Arito etal.. 1997).
11 As humans do not appear to exhibit decreased HR, MAP, and Tco with routine
12 environmental exposures to O3, caution must be used in extrapolating the results of these
13 animal studies to humans (Section 6.3.1).
14 Other studies have shown that O3 can increase BP in animal models. Rats exposed to
15 0.6 ppm O3 for 33 days had increased systolic pressure and HR (Revis etal.. 1981).
16 Increased BP triggers the release of atrial natriuretic factor (ANF), which has been found
17 in increased levels in the heart, lungs, and circulation of O3 exposed (0.5 ppm) rats
18 (Vesely et al.. 1994a. b, c). High concentration O3 exposure (1.0 ppm) has also been
19 found to lead to heart and lung edema (Friedman et al.. 1983). which could be the result
20 of increased ANF levels. Thus, O3 may increase blood pressure and HR, leading to
21 increased ANF and tissue edema.
22 The toxicological studies that have examined the effect of O3 on the cardiovascular
23 system demonstrate O3-induced responses, but it remains unclear if the mechanism is
24 through a reflex response or due to O3 reaction products, which have been sparsely
25 studied. Oxysterols derived from cholesterol ozonation, such as (3-epoxide and 5(3,6(3-
26 epoxycholesterol (and its metabolite cholestan-6-oxo-3,5-diol), have been implicated in
27 inflammation associated with cardiovascular disease (Pulfer et al.. 2005; Pulfer and
28 Murphy. 2004). Two other cholesterol ozonolysis products, atheronal-A and -B (e.g.
29 cholesterol secoaldehyde), have been found in human atherosclerotic plaques and shown
30 in vitro to induce foam cell formation and induce cardiomyocyte apoptosis and necrosis
31 (Sathishkumar et al.. 2005; Wentworth et al.. 2003); however, these products have not
32 been found in the lung compartment or systemically after O3 exposure. The ability to
33 form these cholesterol ozonation products in the circulation in the absence of O3 exposure
34 complicates their implication in O3 induced cardiovascular disease.
3 5 Although it has been proposed that O3 reaction products released after the interaction of
36 O3 with ELF constituents (See Section 5.1.2 on O3 interaction with ELF) are responsible
37 for systemic effects, it is not known whether they gain access to the vascular space.
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1 Alternatively, extrapulmonary release of diffusible mediators, such as cytokines or
2 endothelins, may initiate or propagate inflammatory responses in the vascular or systemic
3 compartments (Cole and Freeman. 2009) (Section 5.1.9.1). Ozone reacts within the lung
4 to amplify ROS production, induce pulmonary inflammation, and activate inflammatory
5 cells, resulting in a cascading proinflammatory state and extrapulmonary release of
6 diffusible mediators that could lead to cardiovascular injury.
6.3.3.2 Recent Cardiovascular Toxicology Studies
7 According to recent short-term O3 exposure animal toxicology studies, O3 plays a role in
8 inducing vascular oxidative stress and proinflammatory mediators, altering HR and HRV,
9 and regulating the pulmonary endothelin system (study details are provided in Table 6-
10 39). A number of these effects were variable between strains examined, suggesting a
11 genetic component to development of O3 induced cardiovascular effects. Further, new
12 studies provide evidence that extended O3 exposure enhances susceptibility to ischemia-
13 reperfusion (I/R) injury and atherosclerotic lesion development. Still, few studies have
14 investigated the role of O3 reaction products in these processes, but more evidence is
15 provided for elevated inflammatory and reduction-oxidation (redox) cascades known to
16 initiate these cardiovascular pathologies.
17 A recent study in young mice and rhesus monkeys examined the effects of short-term O3
18 exposure on a number of cardiovascular endpoints (Chuang et al.. 2009). Mice exposed to
19 O3 for 5 days had increased HR as well as mean and diastolic blood pressure. Increased
20 blood pressure could be explained by the inhibition in endothelial-dependent
21 (acetylcholine) vasorelaxation from decreased bioavailability of aortic nitric oxide (-NO).
22 Ozone caused a decrease in aortic NOX (nitrite and nitrate levels) and a decrease in total,
23 but not phosphorylated, endothelial nitric oxide synthase (eNOS). Ozone also increased
24 vascular oxidative stress in the form of increased aortic and lung lipid peroxidation (F2-
25 isoprostane), increased aortic protein nitration (3-nitrotyrosine), decreased aortic
26 superoxide dismutase (SOD2) protein and activity, and decreased aortic aconitase
27 activity, indicating specific inactivation by O2~ and ONOO". Mitochondrial DNA
28 (mtDNA) damage was also used as a measure of oxidative and nitrative stress in mice
29 and infant rhesus monkeys exposed to O3. Chuang et al. (2009) observed that MtDNA
30 damage accumulated in the lung and aorta of mice after 1 and 5 days of O3 exposure and
31 in the proximal and distal aorta of O3 treated nonhuman primates. Additionally,
32 genetically hyperlipidemic mice exposed to O3 for 8 weeks had increased aortic
33 atherosclerotic lesion area (Section 7.3.1), which may be associated with the short-term
34 exposure changes discussed. Overall, this study suggests that O3 initiates an oxidative
35 environment by increasing O2~ production, which leads to mtDNA damage and -NO
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1 consumption, known to perturb endothelial function (Chuang et al., 2009). Endothelial
2 dysfunction is characteristic of early and advanced atherosclerosis and coincides with
3 impaired vasodilation and blood pressure regulation.
4 Vascular occlusion resulting from atherosclerosis can block blood flow causing ischemia.
5 The restoration of blood flow in the vessel or reperfusion can cause injury to the tissue
6 from subsequent inflammation and oxidative damage. Perepu et al. (2010) observed that
7 O3 exposure enhanced the sensitivity to myocardial I/R injury in rats while increasing
8 oxidative stress levels and pro-inflammatory mediators and decreasing production of anti-
9 inflammatory proteins. Ozone was also found to decrease the left ventricular developed
10 pressure, rate of change of pressure development, and rate of change of pressure decay
11 while increasing left ventricular end diastolic pressure in isolated perfused hearts. In this
12 ex vivo heart model, O3 induced oxidative stress by decreasing SOD enzyme activity and
13 increasing malondialdehyde levels. Ozone also elicited a proinflammatory state which
14 was evident by an increase in TNF-a and a decrease in the anti-inflammatory cytokine
15 IL-10. Perepu et al. (2010) concluded that O3 exposure may result in a greater I/R injury.
Heart Rate and Heart Rate Variability
16 Strain differences in HR and HRV have been observed in response to a 2-h O3
17 pretreatment followed by exposure to carbon black (CB) in mice (C3H/HeJ [HeJ],
18 C57BL/6J [B6], and C3H/HeOuJ [OuJ]) (Hamade and Tankerslev. 2009: Hamade et al..
19 2008). These mice strains were chosen from prior studies on lung inflammatory and
20 hyperpermeability responses to be susceptible (B6 and OuJ) and resistant (HeJ) to O3-
21 induced health effects (Kleeberger et al.. 2000). HR decreased during O3 pre-exposure for
22 all strains, but recovered during the CB exposure (Hamade et al.. 2008). This is contrary
23 to the tachycardia that was reported in 6-week-old B6 mice treated on 1 or 5 days with
24 O3, as described above (Chuang et al.. 2009). Percent change in HRV parameters, SDNN
25 (indicating total HRV) and rMSSD (indicating beat-to-beat HRV), were increased in both
26 C3H mice strains, but not B6 mice, during O3 pre-exposure and recovered during CB
27 exposure when compared to the filtered air group. The two C3H strains differ by a
28 mutation in the Toll-like receptor 4 (TLR4) gene, but these effects did not seem to be
29 related to this mutation since similar responses were observed. Hamade et al. (2008)
30 speculate that the B6 and C3H strains differ in mechanisms of HR response after O3
31 exposure between withdrawal of sympathetic tone and increase of parasympathetic tone;
32 however, no direct evidence for this conclusion was reported. The strain differences
33 observed in HR and HRV suggest that genetic variability affects cardiac responses after
34 acute air pollutant exposures.
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1 Hamade and Tankersley (2009) continued this investigation of gene-environment
2 interactions on cardiopulmonary adaptation of O3 and CB induced changes in HR and
3 HRV using the previously described (Hamade et al., 2008) daily exposure scheme for 3
4 consecutive days. By comparing day-1 interim values it is possible to observe that O3
5 exposure increased SDNN and rMSSD, but decreased HR in all strains. Measures of HR
6 and HRV in B6 and HeJ mice recovered to levels consistent with filtered air treated mice
7 by day 3; however, these responses in OuJ mice remained suppressed. B6 mice had no
8 change in respiratory rate (RR) after O3 treatment, whereas HeJ mice on days 1 and 2 had
9 increased RR and OuJ mice on days 2 and 3 exhibited increased RR. VT did not change
10 with treatment among the strains. Overall, B6 mice were mildly responsive with rapid
11 adaptation, whereas C3 mice were highly responsive with adaptation only in HeJ mice
12 with regards to changes in cardiac and respiratory responses. HR and HRV parameters
13 were not equally correlated with VT and RR between the three mice strains, which
14 suggest that strains vary in the integration of the cardiac and respiratory systems. These
15 complex interactions could help explain variability in interindividual susceptibility to
16 adverse health effects of air pollution.
17 Hamade et al. (2010) expanded their investigation to explore the variation of these strain
18 dependent cardiopulmonary responses with age. As was observed previously, all
19 experimental mouse strains (B6, HeJ, and OuJ) exhibited decreased HR and increased
20 HRV after O3 exposure. Younger O3-exposed mice had a significantly lower HR
21 compared to older exposed mice, indicating an attenuation of the bradycardic effect of O3
22 with age. Younger mice also had a greater increase in rMSSD in HeJ and OuJ strains and
23 SDNN in HeJ mice. Conversely, B6 mice had a slightly greater increase in SDNN in
24 aged mice compared to the young mice. No change was observed in the magnitude of the
25 O3 induced increase of SDNN in OuJ mice or rMSSD in B6 mice. The B6 and HeJ mice
26 genetically vary in respect to the nuclear factor erythroid 2-related factor 2 (Nrf-2). The
27 authors propose that the genetic differences between the mice strains could be altering the
28 formation of ROS, which tends to increase with age, thus modulating the changes in
29 cardiopulmonary physiology after O3 exposure.
30 Strain and age differences in HR and heart function were further investigated in B6 and
31 12981/SvlmJ (129) mice in response to a sequential O3 and filtered air or CB exposure
32 (Tankersley et al., 2010). Young 129 mice showed a decrease in HR after O3 or O3 and
33 CB exposure. This bradycardia was not observed in B6 or older animals in this study,
34 suggesting a possible alteration or adaptation of the autonomic nervous system activity
35 with age. However, these authors did previously report bradycardia in similarly aged
36 young B6 mice (Hamade et al.. 2010; Hamade and Tankersley. 2009; Hamade et al..
37 2008). Ozone exposure in 129 mice also resulted in an increase in left ventricular
3 8 chamber dimensions at end diastole (LVEDD) in young and old mice and a decrease in
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1 left ventricular posterior wall thickness at end systole (PWTES) in older mice. The
2 increase in LVEDD caused a decrease in fractional shortening, which can be used as a
3 rough indicator of left ventricular function. Regression analysis revealed a significant
4 interaction between age and strain on HR and PWTES, which implies that aging affects
5 the HR and function in response to O3 differently between mouse strains.
Effects on Cardiovascular-Related Proteins
6 Increased BP, changes in HRV, and increased atherosclerosis may be related to increases
7 in the vasoconstrictor peptide, endothelin-1 (amino acids 1-21, ET-l[i_2i]). Regulation of
8 the pulmonary endothelin system can be affected in rats by inhalation of PM (0, 5,
9 50 mg/m3, EHC-93) and O3 (Thomson et al.. 2006; Thomson etal.. 2005). Exposure to
10 either O3 (0.8 ppm) or PM increased plasma ET-l^i], ET-3[1.2i], and the ET-1 precursor
11 peptide, bigET-1. Increases in circulating ET-1 [1-21] could be a result of a transient
12 increase in the gene expression of lung preproET-1 and endothelin converting enzyme-1
13 (ECE-1) immediately following inhalation of O3 or PM. These latter gene expression
14 changes (e.g. preproET-1 and ECE-1) were additive with co-exposure to O3 and PM.
15 Conversely, preproET-3 decreased immediately after O3 exposure, suggesting the
16 increase in ET-3[!_21] was not through de novo production. A recent study also found
17 increased ET-1 gene expression in the aorta of O3 exposed rats (Kodavanti et al.. 2011).
18 These rats also exhibited an increase in ETBR after O3 exposure; however, they did not
19 demonstrate increased biomarkers for vascular inflammation, thrombosis, or oxidation.
20 O3 can oxidize protein functional groups and disturb the affected protein. For example,
21 the soluble plasma protein fibrinogen is oxidized by O3 (0.01-0.03 ppm) in vitro, creating
22 fibrinogen and fibrin aggregates, characteristically similar to defective fibrinogen
23 (Rosenfeld et al., 2009; Rozenfeld et al., 2008). In these studies, oxidized fibrinogen
24 retained the ability to form fibrin gels that are involved in coagulation, however the
25 aggregation time increased and the gels were rougher than normal with thicker fibers.
26 Oxidized fibrinogen also developed the ability to self assemble creating fibrinogen
27 aggregates that may play a role in thrombosis. Since O3 does not readily translocate past
28 the ELF and pulmonary epithelium and fibrinogen is primarily a plasma protein, it is
29 uncertain if O3 would have the opportunity to react with plasma fibrinogen. However,
30 fibrinogen can be released from the basolateral face of pulmonary epithelial cells during
31 inflammation, where the deposition of fibrinogen could lead to lung injury (Lawrence
32 and Simpson-Haidaris. 2004).
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Studies on Ozone Reaction Products
1 Although recent toxicological studies have demonstrated O3-induced effects on the
2 cardiovascular system, as concluded in previous O3 AQCDs, it remains unclear if the
3 mechanism is through a reflex response or the result of effects from O3 reaction products
4 (U.S. EPA. 2006b. 1996a). A new study that examined O3 reaction byproducts has shown
5 that cholesterol secoaldehyde (e.g., atheronal A) induces apoptosis in vitro in mouse
6 macrophages (Gao et al., 2009b) and cardiomyocytes (Sathishkumar et al., 2009).
7 Additionally, atheronal-A and -B has been found to induce in vitro macrophage and
8 endothelial cell proinflammatory events involved in the initiation of atherosclerosis
9 (Takeuchi et al.. 2006). These O3 reaction products when complexed with low density
10 lipoprotein upregulate scavenger receptor class A and induce dose-dependent
11 macrophage chemotaxis. Atheronal-A increases expression of the adhesion molecule, E-
12 selectin, in endothelial cells, while atheronal-B induces monocyte differentiation. These
13 events contribute to both monocyte recruitment and foam cell formation in
14 atherosclerotic vessels. It is unknown whether these O3 reaction products gain access to
15 the vascular space from the lungs. Alternative explanations include the extrapulmonary
16 release of diffusible mediators that may initiate or propagate inflammatory responses in
17 the vascular or systemic compartments.
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Table 6-39 Characterization of study details for Section 6.3.3.2a
Study
Model
O3 (ppm) Exposure Duration
Effects
Chuang et al. (2009)
Mice; C57BI/6; M;
6 weeks
0.5
1 or 5 days, 8-h/day
Monkey; rhesus Macaca
mulatta;M; Infant (180
days old)
0.5
5 days, 8-h/day
Increased HR and blood pressure. Initiated an
oxidative environment by increasing vascular
02~ production, which lead to mtDNA damage
. and -NO consumption, known to perturb
endothelial function.
Perepuetal. (2010)
Rat; Sprague-Dawley; 50- 0.8
75 g
28 days, 8-h/day Enhanced the sensitivity to myocardial I/R
injury while increasing oxidative stress and
pro-inflammatory mediators and decreasing
production of anti-inflammatory proteins.
Hamade et al. (2008) Mice; C57BI/6J, C3H/HeJ, 0.6 2-h
and C3H/HeOuJ; M; (subsequent followed by 3 h of CB
CB exposure,
536 ug/m3)
18-20 weeks
Decreased HR. Strain differences observed in
HRV suggest that genetic variability affects
cardiac responses.
Hamade and Tankersley Mice; C57BI/6J, C3H/HeJ, 0.6 3 days, 2-h/day
(2009) and_C3H/HeOuJ; M; (subsequent followed by 3-h of CB
CB exposure,
536 ug/m3)
18-20 weeks
Strains varied in integration of the cardiac and
respiratory systems, implications in
interindividual variability. B6 mice were mildly
responsive with rapid adaptation, whereas C3
mice were highly responsive with adaptation
only in HeJ mice with regards to changes in
cardiac and respiratory responses.
Hamade et al. (2010) Mice; C57BI/6J, C3H/HeJ, 0.6 2-h
andC3H/HeOuJ;M; (subsequent followed by 3-h of CB
5 or 12 mo old CB exposure,
536 ug/m3)
Aged mice exhibited attenuated changes in
cardiopulmonary physiology after 03 exposure.
Genetic differences between mice strains
could be altering formation of ROS, which
tends to increase with age, thus modulating 03
induced effects.
Tankersley et al. (2010) Mice; C57BI/6J,
129S1/SvlmJ;M/F;
5 or 18 mo old
0.6
2-h
(subsequent followed by 3-h of CB
CB exposure,
556 ug/m3)
Significant interaction between age and strain
on HR and PWTES, which implies that aging
affects the HR and function in response to 03
differently between mouse strains.
Thomson et al. (2005) Rat; Fischer-344; M; 200- 0.4 or 0.8
250 g
4-h
Activation of the vasoconstricting ET system.
Increased plasma ET-1 through higher
production and slower clearance.
Thomson et al. (2006)
Kodavanti et al. (2011)
Rat; Fischer-344; M; 200- 0.8
250 g
Rat; Wistar; M; 0.5 or 1.0
10-1 2 weeks
4-h
2 days, 5-h/day
Increased plasma ET-3 not due to de novo
synthesis, unlike ET-1.
No changes to aortic genes of thrombosis,
inflammation or proteolysis, except ET-1 and
ETBR(LOppm).
* Results from previous studies are presented in Table AX5-14 of the 2006 03 AQCD and Table 6-23 of the 1996 03 AQCD.
1
2
3
4
5
6
7
Summary of Toxicological Studies
Overall, animal studies suggest that O3 exposure may disrupt both the -NO and
endothelin systems, which can result in an increase in HR, HRV, and ANF, as is
observed after O3 exposure. Conversely, studies in rodents also exhibit O3 induced
bradycardia, but it is uncertain if this decrease in HR is also observed in humans.
Additionally, O3 may increase oxidative stress and vascular inflammation promoting the
progression of atherosclerosis and leading to increased susceptibility to I/R injury. As O3
reacts quickly with the ELF and does not translocate to the heart and large vessels,
studies suggest that the cardiovascular effects exhibited could be caused by reaction
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September 2011
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1 byproducts of O3 exposure. However, direct evidence of translocation of O3 reaction
2 products to the cardiovascular system has not been demonstrated in vivo. Alternatively,
3 extrapulmonary release of diffusible mediators, such as cytokines or endothelins, may
4 initiate or propagate inflammatory responses in the vascular or systemic compartments
5 leading to the reported cardiovascular pathologies. Further discussion of the modes of
6 action that may lead to cardiovascular effects can be found in Section 5.3.8.
6.3.4 Summary and Causal Determination
7 In past O3 AQCDs the effects of O3 to the cardiovascular system did not receive much
8 attention due to the paucity of information available. However, in recent years,
9 investigation of O3-induced cardiovascular events has advanced. In general, compared
10 with the epidemiologic evidence, the toxicological evidence is more supportive of O3-
11 induced cardiovascular effects. Epidemiologic evidence does not consistently
12 demonstrate a positive relationship between short-term O3 exposure and cardiovascular-
13 related morbidity. However, most epidemiologic studies have not extensively
14 investigated the cardiovascular effects of O3 exposure in susceptible populations, which
15 may further support the toxicological findings. Although the epidemiologic evidence of
16 cardiovascular morbidity is limited, single-city studies reviewed in the 2006 O3 AQCD,
17 recent multicity studies, and the multicontinent APHENA study provide evidence of
18 consistently positive associations between short-term O3 exposure and cardiovascular
19 mortality. However, in contrast with respiratory effects, there is weak coherence between
20 associations for cardiovascular morbidity and mortality. Further, there is no apparent
21 biological mechanism to explain the association observed for short-term O3 exposure
22 with cardiovascular mortality.
23 Animal toxicological studies provide evidence for O3-induced cardiovascular effects,
24 specifically enhanced I/R injury, disrupted NO-induced vascular reactivity, decreased
25 cardiac function, and increased HRV. The observed increase in HRV is supported by a
26 recent controlled human exposure study that also finds increased high frequency HRV,
27 but not altered blood pressure, following O3 exposure. Toxicological studies investigating
28 the role of O3 in heart rate regulation are mixed with both bradycardie and tachycardic
29 responses observed. These changes in cardiac function provide evidence for O3-induced
30 alterations in the autonomic nervous system leading to cardiovascular complications.
31 Epidemiologic studies showing positive association between O3 and arrhythmias confirm
32 the development of autonomic dysfunction following O3 exposure. It is still uncertain
33 how O3 inhalation may cause systemic toxicity; however the cardiovascular effects of O3
34 found in animals correspond to the development and maintenance of an extrapulmonary
3 5 oxidative, proinflammatory environment.
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1 In conclusion, animal toxicological studies provide stronger evidence for O3 exposure
2 leading to cardiovascular morbidity than do epidemiologic studies, among which there is
3 a lack of coherence among endpoints. Based on the relatively strong body of
4 toxicological evidence, and the consistent evidence of an association between O3 and
5 cardiovascular mortality, but weak coherence and biological plausibility for O3-induced
6 cardiovascular morbidity, the generally limited body of evidence is suggestive of a
7 causal relationship between relevant short-term exposures to O3 and
8 cardiovascular effects.
6.4 Central Nervous System Effects
9 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are
10 associated with alterations in neurotransmitters, motor activity, short and long term
11 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have
12 been observed. Reports of headache, dizziness, and irritation of the nose with O3
13 exposure are common complaints in humans, and some behavioral changes in animals
14 may be related to these symptoms rather than indicative of neurotoxicity. Peterson and
15 Andrews (1963) and Tepper et al. (1983) showed that mice would alter their behavior to
16 avoid O3 exposure. Murphy et al. (1964) and Tepper et al. (1982) showed that running -
17 wheel behavior was suppressed, and Tepper et al. (1985) subsequently demonstrated the
18 effects of a 6-h exposure to O3 on the suppression of running-wheel behavior in rats and
19 mice, with the lowest effective concentration being about 0.12 ppm O3 in the rat and
20 about 0.2 ppm in the mouse. The suppression of active behavior by 6 h of exposure to
21 0.12 ppm O3 has recently been confirmed by Martrette et al. (2011) in juvenile female
22 rats, and the suppression of three different active behavior parameters was found to
23 become more pronounced after 15 days of exposure. A table of studies examining the
24 effects of O3 on behavior can be found on p 6-128 of the 1996 O3 AQCD. Generally
25 speaking, transient changes in behavior in rodent models appear to be dependent on a
26 complex interaction of factors such as (1) the type of behavior being measured, with
27 some behaviors increased and others suppressed; (2) the factors motivating that behavior
28 (differences in reinforcement); and (3) the sensitivity of the particular behavior (e.g.,
29 active behaviors are more affected than more sedentary behaviors). Many behavioral
30 changes are likely to result from avoidance of irritation, but more recent studies indicate
31 that O3 also directly affects the CNS.
32 Research in the area of O3-induced neurotoxicity has notably increased over the past few
33 years, with the majority of the evidence coming from toxicological studies that examined
34 the association between O3 exposure, neuropathology, and neurobehavioral effects, and
35 more limited evidence from epidemiologic studies. In an epidemiologic study conducted
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1 by Chen and Schwartz (2009), data from the NHANES III cohort was utilized to study
2 the relationship between long-term O3 exposure (mean annual O3 concentration of
3 26.5 ppb) and neurobehavioral effects among adults aged 20-59 years. The authors
4 observed an association between annual exposure to O3 and tests measuring coding
5 ability and attention/short-term memory. Each 10-ppb increase in annual O3 levels
6 corresponded to an aging-related cognitive performance decline of 3.5 years for coding
7 ability and 5.3 years for attention/short-term memory. These associations persisted in
8 both crude and adjusted models. There was no association between annual O3
9 concentrations and reaction time tests. The authors conclude that overall there is a
10 positive association between O3 exposure and reduced performance on neurobehavioral
11 tests. Although Chen and Schwartz (2009) is a long-term exposure study, it is included in
12 this section because it is the first epidemiologic study to demonstrate that exposure to
13 ambient O3 is associated with decrements in neurocognitive tests related to memory and
14 attention in humans. This epidemiologic evidence of an effect on the CNS due to
15 exposure to ambient concentrations of O3 is coherent with animal studies demonstrating
16 that exposure to O3 can produce a variety of CNS effects including behavioral deficits,
17 morphological changes, and oxidative stress in the brains of rodents. In these rodent
18 studies, interestingly, CNS effects were reported at O3 concentrations that were generally
19 lower than those concentrations commonly observed to produce pulmonary or cardiac
20 effects in rats.
21 A number of new studies demonstrate various perturbations in neurologic function or
22 histology, including changes similar to those observed with Parkinson's and Alzheimer's
23 disease pathologies occurring in similar regions of the brain (Table 6-40). Many of these
24 include exposure durations ranging from short-term to long-term, and as such are
25 discussed here and in Chapter 7 with emphasis on the effects resulting from exposure
26 durations relevant to the respective chapter. Several studies assess short- and long-term
27 memory acquisition via passive avoidance behavioral testing and find decrements in test
28 performance after O3 exposure, consistent with the aforementioned observation made in
29 humans by Chen and Schwartz (2009). Impairment of long-term memory has been
30 previously described in rats exposed to 0.2 ppm O3 for 4 h (Rivas-Arancibia et al.. 1998)
31 and in other studies of 4-hour exposures at concentrations of 0.7 to 1 ppm (Dorado-
32 Martinez etal.. 2001: Rivas-Arancibia et al.. 2000: Avila-Costa et al.. 1999). More
33 recently, statistically significant decreases in both short and long-term memory were
34 observed in rats after 15 days of exposure to 0.25 ppm O3 (Rivas-Arancibia et al.. 2010).
35 The central nervous system is very sensitive to oxidative stress, due in part to its high
36 content of polyunsaturated fatty acids, high rate of oxygen consumption, and low
37 antioxidant enzyme capacity. Oxidative stress has been identified as one of the
3 8 pathophysiological mechanisms underlying neurodegenerative disorders such as
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1 Parkinson's and Alzheimer's disease, among others (Simonian and Coyle. 1996). It is
2 also believed to play a role in altering hippocampal function, which causes cognitive
3 deficits with aging (Vanguilder and Freeman. 2011). A particularly common finding in
4 studies of O3-exposed rats is lipid peroxidation in the brain, especially in the
5 hippocampus, which is important for higher cognitive function including contextual
6 memory acquisition. Performance in passive avoidance learning tests is impaired when
7 the hippocampus is injured, and the observed behavioral effects are well correlated with
8 histological and biochemical changes in the hippocampus, including reduction in spine
9 density in the pyramidal neurons (Avila-Costa et al.. 1999), lipoperoxidation (Rivas-
10 Arancibia et al. 2010; Dorado-Martinez et al.. 2001). progressive neurodegeneration, and
11 activated and phagocytic microglia (Rivas-Arancibia et al.. 2010). The hippocampus is
12 also one of the main regions affected by age-related neurodegenerative diseases,
13 including Alzheimer's disease, and it may be more sensitive to oxidative damage in aged
14 rats. In a study of young (47 days) and aged (900 days) rats exposed to 1 ppm O3 for 4 h,
15 O3-induced lipid peroxidation occurred to a greater extent in the striatum of young rats,
16 whereas it was highest in the hippocampus in aged rats (Rivas-Arancibia et al.. 2000).
17 Martinez-Canabal et al. (2008) showed exposure of rats to 0.25 ppm, 4h/day, for 7, 15, or
18 30 days increased lipoperoxides in the hippocampus. This effect was observed at day 7
19 and continued to increase with time, indicating cumulative oxidative damage. O3-induced
20 changes in lipid peroxidation, neuronal death, and COX-2 positive cells in the
21 hippocampus could be significantly inhibited by daily treatment with growth hormone
22 (GH), which declines with age in most species. The protective effect of GH on -induced
23 oxidative stress was greatest at 15 days of exposure and was non-significant at day 30.
24 Consistent with these findings, lipid peroxidation in the hippocampus of rats was
25 observed to increase significantly after a 30-day exposure to 0.25 ppm , but not after a
26 single 4-h exposure to the same concentration (Mokoena et al.. 2010). However, 4 hours
27 of exposure was sufficient to cause significant increases in lipid peroxidation when the
28 concentration was increased to 0.7 ppm, and another study observed lipid peroxidation
29 after a 4-h exposure to 0.4 ppm (Dorado-Martinez et al.. 2001).
30 Other commonly affected areas of the brain include the striatum, substantia nigra,
31 cerebellum, olfactory bulb, and frontal/prefrontal cortex. The striatum and substantia
32 nigra are particularly sensitive to oxidative stress because the metabolism of dopamine,
33 central to their function, is an oxidative process perturbed by redox imbalance. Oxidative
34 stress has been implicated in the premature death of substantia nigra dopamine neurons in
35 Parkinson's disease. Angoa-Perez et al. (2006) have shown progressive lipoperoxidation
36 in the substantia nigra and a decrease in nigral dopamine neurons in ovariectomized
37 female rats exposed to 0.25 ppm O3, 4h/day, for 7, 15, or 30 days. Estradiol, an
38 antioxidant, attenuated O3-induced oxidative stress and nigral neuronal death, and the
39 authors note that in humans, estrogen therapy can ameliorate symptoms of Parkinson's
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1 disease, which is more prevalent in men. Progressive oxidative stress has also been
2 observed in the striatum and substantia nigra of rats after 15 and 30 days of exposure to
3 0.25 ppm O3 for 4 h/day, along with a loss of dopaminergic neurons from the substantia
4 nigra (Pereyra-Munoz et al.. 2006). Decreases in motor activity were also observed at 15
5 and 30 days of exposure, consistent with other reports (Martrette et al., 2011; Dorado-
6 Martinez et al.. 2001). Using a similar O3 exposure protocol, Santiago-Lopez and
7 colleagues (2010) also observed a progressive loss of dopaminergic neurons within the
8 substantia nigra, accompanied by alterations in the morphology of remaining cells and an
9 increase in p53 levels and nuclear translocation.
10 The olfactory bulb also undergoes oxidative damage in O3 exposed animals, in some
11 cases altering olfactory-dependent behavior. Lipid peroxidation was observed in the
12 olfactory bulbs of ovariectomized female rats exposed to 0.25 ppm O3 (4 h/day) for 30 or
13 60 days (Guevara-Guzman et al.. 2009). O3 also induced decrements in a selective
14 olfactory recognition memory test, and the authors note that early deficits in odor
15 perception and memory are components of human neurodegenerative diseases. The
16 decrements in olfactory memory were not due to damaged olfactory perception based on
17 other tests. However, deficits in olfactory perception emerged with longer exposures
18 (discussed in Chapter 7). As with the study by Angoa-Perez et al. (2006) described
19 above, a protective effect for estradiol was demonstrated for both lipid peroxidation and
20 olfactory memory defects. The role of oxidative stress in memory deficits and associated
21 morphological changes has also been demonstrated via attenuation by other antioxidants
22 as well, such as a-tocopherol (Guerrero et al.. 1999) and taurine (Rivas-Arancibia et al..
23 2000).
24 It is unclear how persistent these effects might be. One study of acute exposure, using
25 1 ppm O3 for 4 hours, observed morphological changes in the olfactory bulb of rats at
26 2 hours, and 1 and 10 days, but not 15 days, after exposure (Colin-Barenque et al., 2005).
27 Other acute studies also report changes in the CNS. Lipid peroxidation was observed in
28 multiple regions of the brain after a 1- to 9-h exposure to 1 ppm O3 (Escalante-Membrillo
29 et al.. 2005). Ozone has also been shown to alter gene expression of endothelin-1
30 (pituitary) and inducible nitric oxide synthase (cerebral hemisphere) after a single 4-h
31 exposure to 0.8 ppm O3, indicating potential cerebrovascular effects. This concentration-
32 dependent effect was not observed at 0.4 ppm O3 (Thomson et al.. 2007). Vascular
33 endothelial growth factor was upregulated in astroglial cells in the central respiratory
34 areas of the brain of rats exposed to 0.5 ppm O3 for 3 hours (Araneda et al., 2008). The
35 persistence of CNS changes after a single exposure was also examined and the increase in
36 vascular endothelial growth factor was present after a short (3 hours) recovery period.
37 Thus, there is evidence that O3-induced CNS effects are both concentration- and time-
3 8 dependent.
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1 Because O3 can produce a disruption of the sleep-wake cycle (U.S. EPA. 2006b). Alfaro-
2 Rodriguez et al. (2005) examined whether acetylcholine in a region of the brain involved
3 in sleep regulation was altered by O3. After a 24-h exposure to 0.5 ppm O3, the
4 acetylcholine concentration in the medial preoptic area was decreased by 58% and
5 strongly correlated with a disruption in paradoxical sleep. Such behavioral-biochemical
6 effects of O3 are confirmed by a number of studies which have demonstrated
7 morphological and biochemical changes in rats.
8 CNS effects have also been demonstrated in newborn and adult rats whose only exposure
9 to O3 occurred in utero. Several neurotransmitters were assessed in male offspring of
10 dams exposed to 1 ppm O3 during the entire pregnancy (Gonzalez-Pina et al., 2008). The
11 data showed that catecholamine neurotransmitters were affected to a greater degree than
12 indole-amine neurotransmitters in the cerebellum. CNS changes, including behavioral,
13 cellular, and biochemical effects, have also been observed after in utero exposure to
14 0.5 ppm O3 for 12 h/day from gestational days 5-20 (Boussouar et al., 2009). Tyrosine
15 hydroxylase labeling in the nucleus tractus solatarius was increased after in utero
16 exposure to O3 whereas Fos protein labeling did not change. When these offspring were
17 challenged by immobilization stress, neuroplasticity pathways, which were activated in
18 air-exposed offspring, were inhibited in O3-exposed offspring. Although an O3 exposure
19 concentration-response was not studied in these two in utero studies, it has been
20 examined in one study. Santucci et al. (2006) investigated behavioral effects and gene
21 expression after in utero exposure of mice to as little as 0.3 ppm O3. Increased
22 defensive/submissive behavior and reduced social investigation were observed in both the
23 0.3 and 0.6 ppm O3 groups. Changes in gene expression of brain-derived neurotrophic
24 factor (BDNF, increased in striatum) and nerve growth factor (NGF, decreased in
25 hippocampus) accompanied these behavioral changes. Thus, these three studies
26 demonstrate that CNS effects can occur as a result of in utero exposure to O3, and
27 although the mode of action of these effects is not known, it has been suggested that
28 circulating lipid peroxidation products may play a role (Boussouar et al.. 2009).
29 Importantly, these CNS effects occurred in rodent models after in utero only exposure to
3 0 relevant concentrations of O3.
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Table 6-40 Central Nervous System and Behavioral Effects of Short-term O3
Exposure in Rats
Study
Martretteetal. (2011)
Angoa-Perez et al. (2006)
Guevara-Guzman et al. (2009)
Martinez-Canabal et al. (2008)
Pereyra-Munoz et al. (2006)
Rivas-Arancibia et al. (2010)
Santiago-Lopez et al. (2010)
Thomson et al. (2007)
Model , °3 .
(ppm)
Rat; Wistar; F; 0.12
Weight: 152g;
7 weeks old
Rat; Wistar; F; Weight: 0.25
300g; ovariectomized
Rat; Wistar; F; 264g; 0.25
ovariectomized
Rat; Wistar; M; 0.25
Weight: 300g
Rat; Wistar; M; 250- 0.25
300g
Rat; Wistar; M; 250- 0.25
300g
Rat; Wistar; M; 250- 0.25
300g
Rat; Fischer-344; M; 0.4; 0.8
200-250g
Exposure
Duration
1-15days,6h/day
7 to 60 days, 4-
h/day, 5 days/wk
30 and 60 days,
4h/day
7 to 30 days, 4-h/day
15 and 30 days, 4-h/
day
15 to 90 days, 4-h/
day
15, 30, and 60 days,
4-h/day
4-h; assays at 0 and
24 h post exposure
Effects
Significant decrease in rearing, locomotor activity,
and jumping activity at day 1 , with a further
decrease in these activities by day 15.
Progressive lipid peroxidation and loss of tyrosine
hydrolase-immunopositive neurons in the
substantia nigra starting at 7 days.
Estradiol treatment protected against lipid
peroxidation and decreases in estrogen receptors
and dopamine p-hydroxylase in olfactory bulbs
along with deficits in olfactory recognition
memory.
Growth hormone inhibited 03-induced increases
in lipoperoxidation and COX-2 positive cells in the
hippocampus.
Decreased motor activity, increased lipid
peroxidation, altered morphology, and loss of
dopamine neurons in substantia nigra and
striatum, increased expression of DARPP-32,
iNOS, and SOD.
Ozone produced significant increases in lipid
peroxidation in the hippocampus, and altered the
number of p53 positive immunoreactive cells,
activated and phagocytic microglia cells, GFAP
immunoreactive cells, and doublecortine cells,
and short- and long-term memory-retention
latency.
Progressive loss of dopamine reactivity in the
substantia nigra, along with morphological
changes. Increased p53 levels and nuclear
translocation.
At 0.8 ppm, 03 produced rapid perturbations in
the ET-NO pathway gene expression in the brain.
Ozone induced a small but significant time- and
concentration-dependent increase in prepro-
endothelin-1 mRNA levels in the cerebral
hemisphere and pituitary, whereas TNFa and
iNOS mRNA levels were decreased at 0 hrs and
unchanged or increased, respectively, at 24 h.
Alfaro-Rodriguez and Gonzalez-
Pina (2005)
Rat; Wistar; M; 292g 0.5
24-h
During the light phase, 03 caused a significant
decrease in paradoxical sleep accompanied by a
significant decrease in Ach levels in the
hypothalamic medial preoptic area. The same
effects occurred during the dark phase exposure
to 03 in addition to a significant increase in slow-
wave sleep and decrease in wakefulness.
Araneda et al. (2008)
Boussouar et al. (2009)
Rats; Sprague- 0.5
Dawley; M; 280-320g
Rat; Sprague-Dawley; 0.5
M; adult offspring of
prenatally exposed
dams; 403-41 4g
3-h (measurements
taken at 0 h and 3 h
after exposure)
From embryonic day
E5toE20for12-
h/day; immobilization
stress
Ozone upregulated VEGF in astroglial cells
located in the respiratory center of the brain.
VEGF co-located with IL-6 and TNF in cells near
blood vessel walls, and blood vessel area was
markedly increased.
Prenatal 03 exposure had a long term impact on
the nucleus tractus solitarius of adult rats, as
revealed during immobilization stress.
Soulage et al. (2004)
Rat; Sprague-Dawley;
M;Approx. 7 weeks
old
0.7
5-h
Ozone produced differential effects on peripheral
and central components of the sympatho-adrenal
system. While catecholamine biosynthesis was
increased in portions of the brain, the
catecholamine turnover rate was significantly
increased in the heart and cerebral cortex and
inhibited in the lung and striatum.
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September 2011
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Study
Guzman et al. (2006: 2005)
Colin-Barenque et al. (2005)
Escalante-Membrillo etal. (2005)
Gonzalez-Pina et al. (2008)
Model
Rat; Wistar; M; 21
days old; well-
nourished and
malnourished groups
Rats; Wistar; M; 250-
300g
Rats; Wistar; M; 280-
320g
Rat; Wistar; M;
Os Exposure
(ppm) Duration
0.75 15 successive days
for 4-h/day
1 .0 4-h; assays at 2-h,
24-h, 10 days, and
15 days after
exposure
1.0 1-,3-,6-,or9-h
1 12-h/day,21daysof
gestation; assays at
0, 5,& 10 days
postnatal
Effects
A significant decrease in body weight was
observed in both well nourished (WN) and
malnourished (MN) rats after 03 exposure.
Localized ATPase, TEARS, and GSH levels
changed in response to ozone in certain brain
areas and the ozone-induced changes were
dependent on nutritional condition.
A significant loss of dendritic spines in granule
cells of the olfactory bulb occurred at 2 hrs to 10
days after exposure. Cytological and
ultrastructural changes returned towards normal
morphology by 15 days.
Significant increases in TEARS occurred in
hypothalamus, cortex, striatum, midbrain,
thalamus, and pons. Partial but significant
recovery was observed by 3 h after the 9 h
exposure.
Prenatal 03 exposure produced significant
decreases in cerebellar monoamine but not
indolamine. content at 0 and 5 days after birth
with a partial recovery by 10 d. 5-hydroxy-indole-
acetic acid levels were significantly increased at
10 days.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
6.4.1 Neuroendocrine Effects
According to the 2006 O3 AQCD, early studies suggested an interaction of O3 with the
pituitary-thyroid-adrenal axis, because thyroidectomy, hypophysectomy, and
adrenalectomy protected against the lethal effects of O3. Concentrations of 0.7-1.0 ppm
O3 for a 1-day exposure in male rats caused changes in the parathyroid, thymic atrophy,
decreased serum levels of thyroid hormones and protein binding, and increased prolactin.
Increased toxicity to O3 was reported in hyperthyroid rats and T3 supplementation was
shown to increase metabolic rate and pulmonary injury in the lungs of O3-treated animals.
The mechanisms by which O3 affects neuroendocrine function are not well understood,
but previous work suggests that high ambient levels of O3 can produce marked neural
disturbances in structures involved in the integration of chemosensory inputs, arousal,
and motor control, effects that may be responsible for some of the behavioral effects seen
with O3 exposure. A more recent study exposing immature female rats to 0.12 ppm O3
demonstrated significantly increased serum levels of the thyroid hormone free T3 after 15
days of exposure, whereas free T4 was unchanged (Martrette et al., 2011). These results
are in contrast to those previously presented whereby 1 ppm O3 for 1 day significantly
decreased T3 and T4 (demons and Garcia. 1980), although comparisons are made
difficult by highly disparate exposure regimens along with sex differences. Martrette et
al.(2011) also demonstrated significantly increased corticosterone levels after 15 days of
exposure, suggesting a stress related response.
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6.4.2 Summary and Causal Determination
1 In rodents, O3 exposure has been shown to cause physicochemical changes in the brain
2 indicative of oxidative stress and inflammation. Newer toxicological studies add to earlier
3 evidence that acute exposures to O3 can produce a range of effects on the central nervous
4 system and behavior. Previously observed effects, including neurodegeneration,
5 alterations in neurotransmitters, short and long term memory, and sleep patterns, have
6 been further supported by recent studies. In instances where pathology and behavior are
7 both examined, animals exhibit decrements in behaviors tied to the brain regions or
8 chemicals found to be affected or damaged. For example, damage in the hippocampus,
9 which is important for memory acquisition, was correlated with impaired performance in
10 tests designed to assess memory. Thus the brain is functionally affected by O3 exposure.
11 The single epidemiologic study conducted showed an association between O3 exposure
12 and memory deficits in humans as well, albeit on a long-term exposure basis. Notably,
13 exposure to O3 levels as low as 0.25 ppm for 7 days has resulted in progressive
14 neurodegeneration and deficits in both short and long-term memory in rodents.
15 Examination of changes in the brain at lower exposure concentrations or at 0.25 ppm for
16 shorter durations has not been reported, but 0.12 ppm O3 has been shown to alter
17 behavior. It is possible that some behavioral changes may reflect avoidance of irritation
18 as opposed to functional changes in brain morphology or chemistry, but in many cases
19 functional changes are related to oxidative stress and damage. In some instances, changes
20 were dependent on the nutritional status of the rats (high versus low protein diet). For
21 example, O3 produced an increase in glutathione in the brains of rats fed the high protein
22 diet but decreases in glutathione in rats fed low protein chow (Calderon Guzman et al.,
23 2006). The hippocampus, one of the main regions affected by age-related
24 neurodegenerative diseases, appears to be more sensitive to oxidative damage in aged rats
25 (Rivas-Arancibia et al., 2000). and growth hormone, which declines with age in most
26 species, may be protective (Martinez-Canabal and Angora-Perez. 2008). Developing
27 animals may also be sensitive, as changes in the CNS, including biochemical, cellular,
28 and behavioral effects, have been observed in juvenile and adult animals whose sole
29 exposure occurred in utero, at levels as a low as 0.3 ppm. A number of studies
30 demonstrate ozone-induced changes that are also observed in human neurodegenerative
31 disorders such as Alzheimer's and Parkinson's disease, including signs of oxidative
32 stress, loss of neurons/neuronal death, reductions in dopamine levels, increased COX-2
33 expression, and increases in activated microglia in important regions of the brain
34 (hippocampus, substantianigra).
35 Although evidence from epidemiologic and controlled human exposure studies is lacking,
36 the toxicological evidence for ozone's impact on the brain and behavior is strong, and at
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1 least is suggestive of a causal relationship between O3 exposure and effects on the
2 central nervous system.
6.5 Effects on Other Organ Systems
6.5.1 Effects on the Liver and Xenobiotic Metabolism
3 Early investigations of the effects of O3 on the liver centered on xenobiotic metabolism,
4 and the prolongation of drug-induced sleeping time, which was observed at 0.1 ppm O3
5 (Graham et al., 1981). In some species, only adults and especially females were affected.
6 In rats, high (1.0-2.0 ppm for 3 hours) acute O3 exposures caused increased production of
7 NO by hepatocytes and enhanced protein synthesis (Laskin et al., 1996; Laskin et al.,
8 1994). Except for the earlier work on xenobiotic metabolism, the responses occurred only
9 after very high acute O3 exposures. One study, conducted at 1 ppm O3 exposure, has been
10 identified (Last et al.. 2005) in which alterations in gene expression underlying O3-
11 induced cachexia and downregulation of xenobiotic metabolism were examined. A
12 number of the down-regulated genes are known to be interferon (IFN) dependent,
13 suggesting a role for circulating IFN. A more recent study by Aibo et al. (2010)
14 demonstrates exacerbation of acetaminophen-induced liver injury in mice after a single
15 6-h exposure to 0.25 or 0.5 ppm O3. Data indicate that O3 may worsen drug-induced liver
16 injury by inhibiting hepatic repair. The O3-associated effects shown in the liver are
17 thought to be mediated by inflammatory cytokines or other cytotoxic mediators released
18 by activated macrophages or other cells in the lungs (Laskin and Laskin. 2001; Laskin et
19 al., 1998; Vincent et al.. 1996b). Recently, increased peroxidated lipids were detected in
20 the plasma of O3 exposed animals (Santiago-Lopez et al.. 2010).
21 In summary, mediators generated by O3 exposure may cause effects on the liver in
22 laboratory rodents. Ozone exposures as low as 0.1 ppm have been shown to affect drug-
23 induced sleeping time, and exposure to 0.25 ppm can exacerbate liver injury induced by a
24 common analgesic. However, very few studies at relevant concentrations have been
25 conducted, and no data from controlled human exposure or epidemiologic studies are
26 currently available. Therefore the collective evidence is inadequate to determine if a
27 causal relationship exists between short-term O3 exposure and effects on the liver
28 and metabolism.
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6.5.2 Effects on Cutaneous and Ocular Tissues
1 In addition to the lungs, the skin is highly exposed to O3 and contains O3 reactive targets
2 (polyunsaturated fatty acids) that can produce lipid peroxides. The 2006 O3 AQCD
3 reported that although there is evidence of oxidative stress at near ambient O3
4 concentrations, skin and eyes are only affected at high concentrations (greater than
5 1-5 ppm). Ozone exposure (0.8 ppm for 7 days) induces oxidative stress in the skin of
6 hairless mice, along with proinflammatory cytokines (Valacchi et al., 2009). A recent
7 study demonstrated that 0.25 ppm O3 differentially alters expression of
8 metalloproteinases in the skin of young and aged mice, indicating age-related
9 susceptibility to oxidative stress (Fortino et al.. 2007). In young mice, healing of skin
10 wounds is not significantly affected by O3 exposure (Lim et al.. 2006). However,
11 exposure to 0.5 ppm O3 for 6 h/day significantly delays wound closure in aged mice. As
12 with effects on the liver described above, the effects of O3 on the skin and eyes have not
13 been widely studied, and information from controlled human exposure or epidemiologic
14 studies is not currently available. Therefore the collective evidence is inadequate to
15 determine if a causal relationship exists between short-term O3 exposure and
16 effects on cutaneous and ocular tissues.
6.6 Mortality
6.6.1 Summary of Findings from 2006 Ozone AQCD
17 The 2006 O3 AQCD reviewed a large number of time-series studies consisting of single-
18 and multicity studies, and meta-analyses. In the large U.S. multicity studies that
19 examined all-year data, summary effect estimates corresponding to single-day lags
20 ranged from a 0.5-1% increase in all-cause (nonaccidental) mortality per the standardized
21 unit increase1 in O3. The association between short-term O3 exposure and mortality was
22 substantiated by a collection of meta-analyses and international multicity studies. The
23 studies evaluated found some evidence for heterogeneity in O3 mortality risk estimates
24 across cities and studies. Studies that conducted seasonal analyses, although more limited
25 in number, reported larger O3 mortality risk estimates during the warm or summer
26 season. Overall, the 2006 O3 AQCD identified robust associations between various
27 measures of daily ambient O3 concentrations and all-cause mortality, with additional
28 evidence for associations with cardiovascular mortality, which could not be readily
29 explained by confounding due to time, weather, or copollutants. However, it was noted
1 In the 2006 O3 AQCD and throughout this document to compare across studies that used the same exposure metric, effect
estimates were standardized to 40 ppb for 1-h maximum, 30 ppb for 8-h maximum, and 20 ppb for 24-h average O3 concentrations.
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1 that multiple uncertainties remain regarding the O3-mortality relationship including: the
2 extent of residual confounding by copollutants; factors that modify the O3-mortality
3 association; the appropriate lag structure for identifying O3-mortality effects (e.g., single-
4 day lags versus distributed lag model); the shape of the O3-mortality C-R function and
5 whether a threshold exists; and the identification of susceptible populations. Collectively,
6 the 2006 O3 AQCD concluded that "the overall body of evidence is highly suggestive that
7 O3 directly or indirectly contributes to non-accidental and cardiopulmonary-related
8 mortality."
6.6.2 Associations of Mortality and Short-Term Ozone Exposure
9 The recent literature that examined the association between short-term O3 exposure and
10 mortality further confirmed the associations reported in the 2006 O3 AQCD. New
11 multicontinent and multicity studies reported consistent positive associations between
12 short-term O3 exposure and all-cause mortality in all-year analyses, with additional
13 evidence for larger mortality risk estimates during the warm or summer months (Figure
14 6-27; Table 6-41). These associations were reported across a range of ambient O3
15 concentrations that were in some cases quite low (Table 6-42).
Draft - Do Not Cite or Quote 6-194 September 2011
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Study
Gryparisetal. (2004;57276)
Bell etal. (2007; 93256)
Schwartz (2005; 57333)
Bell and Dominici (2008; 193828)
Bell etal. (2004; 94417)a
Levy et al. (2005; 74347)a
Katsouyanni et al. (2009; 199899)
Bell etal. (2005; 74345 )a
Ito etal. (2005; 74346)a
Wonget al. (2010; 732535)
Katsouyanni et al. (2009; 199899)
Cakmak etal. (2011, 699135)
Katsouyanni et al. (2009; 199899
Katsouyanni et al. (2009; 199899 b
Samolietal.(2009;195855)
Bell etal. (2004; 94417)a
Schwartz (2005; 57333)
Zanobetti and Schwartz (2008; 195755
Zanobetti and Schwartz (2008; 101596
Franklin and Schwartz (2008; 156448)
Gryparisetal. (2004;57276)
Medina-Ramon and Schwartz (2008)
Katsouyanni et al. (2009; 199899)
Bell etal. (2005; 74345 )a
Katsouyanni et al. (2009; 199899
Katsouyanni et al. (2009; 199899 b
Levy etal. (2005; 74 347)a
Ito etal. (2005; 74346)a
Katsouyanni et al. (2009; 199899)
Stafoggia et al. (2010; 625034)
Location
APHEA2 (23 cities)
98 U.S. communities
14 U.S. cities
98 U.S. communities
95 U.S. communities
U.S. and Non-U.S.
APHENA-Europe
U.S. and Non-U.S.
U.S. and Non-U.S.
PAPA (4 cities
APHENA-y.S.
7 Chilean cities
APHENA-Canada
APHENA-Canada
21 European cities
95 U.S. communities
48 U.S. cities
48 U.S. cities
18 U.S. communities
APHEA2 (21 cities)
48 U.S. cities
APHENA-Europe
U.S. and Non-U.S.
APHENA-Canada
APHENA-Canada
U.S. and Non-U.S.
U.S. and Non-U.S.
APHENA-U.S.
10 Italian cities
Lag
0-1
0-1
0°6
0-6
DL(0-2)
0-1
DL 0-2
All-Year
DL
DL
0-6
DL 0-2
0-2
0-1
0Q6
0
0-3
0
0-1
0-2
DL(0-2)
DLjO-2)
DL(0-2)
DL(0-2)
DL(0-5)
Summer
11
% Increase
Effect estimates are for a 40 ppb increase in 1-h max, 30 ppb increase in 8-h max, and 20 ppb increase in 24-h avg ozone
concentrations. An "a" represent multicity studies and meta-analyses from the 2006 ozone AQCD. Bell et al. (2005). Ito et al. (2005).
and Levy et al. (2005) used a range of lag days in the meta-analysis: Lag 0, 1, 2, or average 0-1 or 1-2; single-day lags from 0 to 3;
and lag 0 and 1-2; respectively. A"b" represents risk estimates from APHENA-Canada standardized to an approximate IQR of 5.1
ppb for a 1-h max increase in ozone concentrations (see explanation in Section 6.2.7.2).
Figure 6-27 Summary of mortality risk estimates for short-term ozone exposure
and all-cause (nonaccidental) mortality from all-year and summer
season analyses.
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Table 6-41 Corresponding effect estimates for Figure 6-27
Study
Location
Lag
Avg Time
% Increase (95% Cl)
All-year
Gryparis et al. (2004)
Bell et al. (2007)
Schwartz (2005a)
Bell and Dominici (2008)
Bell et al. (2004)a
Lew et al. (2005)a
Katsouyanni et al. (2009)
Bell et al. (2005)a
Ito et al. (2005)a
Wonaetal. (2010)
Katsouyanni et al. (2009)
Cakmaketal. (2011)
Katsouyanni et al. (2009)
Katsouyanni et al. (2009)b
APHEA2 (23 cities)
98 U.S. communities
14 U.S. cities
98 U.S. communities
95 U.S. communities
U.S. and Non-U.S.
APHENA-Europe
U.S. and Non-U.S.
U.S. and Non-U.S.
PAPA (4 cities)
APHENA-U.S.
7 Chilean cities
APHENA-Canada
APHENA-Canada
0-1
0-1
0
0-6
0-6
—
DL(0-2)
—
—
0-1
DL(0-2)
DL(0-6)
DL(0-2)
DL(0-2)
1-hmax
24-h avg
1-hmax
24-h avg
24-h avg
24-h avg
1-hmax
24-h avg
24-h avg
8-h avg
1-hmax
8-h max
1-h max
1-h max
0.24 (-0.86, 1.98)
0.64 (0.34, 0.92)
0.76(0.13,1.40)
1.04(0.56,1.55)
1.04(0.54,1.55)
1.64(1.25,2.03)
1 .66 (0.47, 2.94)
1.75(1.10,2.37)
2.20 (0.80, 3.60)
2.26(1.36,3.16)
3.02(1.10,4.89)
3.35(1.07,5.75)
5.87(1.82,9.81)
0.73(0.23,1.20)
Summer
Samolietal. (2009)
Bell et al. (2004)a
Schwartz (2005a)
Zanobetti and Schwartz (2008a)
Zanobetti and Schwartz (2008b)
Franklin and Schwartz (2008)
Gryparis et al. (2004)
Medina-Ramon and Schwartz (2008)
Katsouyanni et al. (2009)
Bell et al. (2005)a
Katsouyanni et al. (2009)
Katsouyanni et al. (2009)
Lew et al. (2005)a
Ito et al. (2005)a
Katsouyanni et al. (2009)
Stafoggiaetal. (2010)
21 European cities
95 U.S. communities
14 U.S. cities
48 U.S. cities
48 U.S. cities
18 U.S. communities
APHEA2 (21 cities)
48 U.S. cities
APHENA-Europe
U.S. and Non-U.S.
APHENA-Canada
APHENA-Canada
U.S. and Non-U.S.
U.S. and Non-U.S.
APHENA-U.S.
10 Italian cities
0-1
0-6
0
0
0-3
0
0-1
0-2
DL(0-2)
—
DL(0-2)
DL(0-2)
—
—
DL(0-2)
DL(0-5)
8-h max
24-h avg
1-h max
8-h max
8-h max
24-h avg
8-h max
8-h max
1-hmax
24-h avg
1-hmax
1-hmax
24-h avg
24-h avg
1-hmax
8-h max
0.66(0.24,1.05)
0.78(0.26,1.30)
1.00(0.30,1.80)
1.51 (1.14,1.87)
1 .60 (0.84, 2.33)
1 .79 (0.90, 2.68)
1 .80 (0.99, 3.06)
1.96(1.14,2.82)
2.38(0.87,3.91)
3.02(1.45,4.63)
3.34(1.26,5.38)
0.42 (0.16, 0.67)
3.38 (2.27, 4.42)
3.50(2.10,4.90)
3.83(1.90,5.79)
9.15(5.41,13.0)
aMulticity studies and meta-analyses from the 2006 03 AQCD. Bell et al. (2005)'. Ito et al. (2005)'. and Levy et al. (2005)' used a range of lag
days in the mete-analysis: Lag 0,1, 2, or average 0-1 or 1-2; Single-day lags from 0-3; and Lag 0 and 1-2; respectively.
bRisk estimates from APHENA-Canada standardized to an approximate IQRof 5.1 ppbfora 1-h max increase in 03 concentrations (see
explanation in Section 6.2.7.2).
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Table 6-42 Range of mean and upper percentile ozone concentrations in
previous and recent multicity studies
1
2
Study
Gryparis et al.
(2004)"
Schwartz (2005a)b
Bell et al. (2004)
Bell et al. (2007)
Bell and Dominici
(2008)
Franklin and
Schwartz (2008)
Katsouyanni et al.
(2009)"'e
Medina-Ramon
and Schwartz
(2008)b
Samolietal.
(2009)b
Stafoggiaetal.
(2010)
Cakmaket al.
(2011)
Wong et al. (2010)
Zanobetti and
Schwartz (2008b)
Zanobetti and
Schwartz (2008a)
Location
Years
23 European 1990-1997
cities (APHEA2)
14 U.S. cities 1986-1993
95 U.S. 1987-2000
communities
(NMMAPS)
98 U.S. 1987-2000
communities
(NMMAPS)
98 U.S. 1987-2000
communities (All year and
(NMMAPS) May-September)
18 U.S. 2000-2005
communities (May-September)
NMMAPS 1987-1996 (Canada and
12 Canadian U.S.) varied by city for
cities Europe
(APHEA2)
48 U.S. cities 1989-2000
(May-September)
21 European 1990-1997
cities (APHEA2) (June-August)
10 Italian cities 2001-2005
(April-September)
7 Chilean cities 1997-2007
PAPA (4 cities) 1999-2003 (Bangkok)
1996-2002 (Hong Kong)
2001-2004
(Shanghai)
2001 -2004 (Wuhan)
48 U.S. cities 1989-2000
(June-August)
48 U.S. cities"
1989-2000
Wnter: December-February)
Spring: March-May)
Summer: June-August)
Autumn: September-
\lovem ber)
Averaging
Time
1-h max
8-h max
1-h max
24-h avg
24-h avg
24-h avg
24-h avg
1-h max
8-h max
8-h max
8-h max
8-h max
8-h avg
8-h max
8-h max
Mean
Concentration (ppb)
Summer:
1-h max: 44-1 17
8-h max: 30-99
Winter:
1-h max: 11 -57
8-h max: 8-49
35.1-60
26.0
26.0°
All year: 26.8
May-September: 30.0
21.4-48.7
U.S.: 13.3-38.4
Canada: 6.7-8.4
Europe:1 8.3-41 .9
16.1-58.8
20.0-62.8
41.2-58.9
59.0-87.6
18.7-43.7
15.1-62.8
Wnter: 16.5
Spring: 41. 6
Summer: 47.8
Autumn: 33.5
Upper Percentile
Concentrations (ppb)
Summer:
1-h max: 62-1 73
8-h max: 57-1 54
Winter:
1-h max: 40-88
8-h max: 25-78
25th: 26.5-52
75th: 46.3-69
NR
NR
Maximum:
All year: 37.3
May-September: 47.2
NR
75th:
U.S.: 21. 0-52.0
Canada: 8.7-1 2.5
Europe: 24.0-65.8
NR
75th: 27.2-74.8
75th: 47.0-71 .6
NR
75th: 38.4 -60.4
Max: 92.1 -131. 8
Max: 34.3-146.2
75th: 19.8-75.9
Max:
Wnter: 40.6
Spring: 91. 4
Summer: 103.0
Autumn: 91.2
a03 concentrations were converted to ppb if the study presented them as ug/m3 by using the conversion factor of 0.51 assuming standard
temperature (25° C) and pressure (1 atm).
bStudy only reported median 03 concentrations.
°Cities with less than 75% observations in a season excluded. As a result, 29 cities examined in winter, 32 in spring, 33 in autumn, and all 48 in
the summer.
dBell et al. (2007)did not report mean 03 concentrations, however, it used a similar dataset as Bell et al. (2004) which consisted of 95 U.S.
communities for 1987-2000. For comparison purposes the 24-h avg 03 concentrations for the 95 communities from Bell et al. (2004) are reported
here.
eStudy did not present air quality data for the summer months.
CV=coefficient of variation
In addition to examining the relationship between short-term O3 exposure and all-cause
mortality, recent studies attempted to address the uncertainties that remained upon the
completion of the 2006 O3 AQCD. As a result, given the robust associations between
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September 2011
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1 short-term O3 exposure and mortality presented across studies in the 2006 O3 AQCD and
2 supported in the new multicity studies, the following sections primarily focus on the
3 examination of previously identified uncertainties in the O3-mortality relationship,
4 specifically: O3 associations with cause-specific mortality, confounding, lag structure
5 (e-g-, multiday effects and mortality displacement), effect modification (i.e., sources of
6 heterogeneity in risk estimates across cities); and the O3-mortality C-R relationship.
7 Focusing specifically on these uncertainties allows for a more detailed characterization of
8 the relationship between short-term O3 exposure and mortality.
6.6.2.1 Confounding
9 Recent epidemiologic studies examined potential confounders of the O3-mortality
10 relationship. These studies specifically focused on whether PM and its constituents or
11 seasonal trends confounded the association between short-term O3 exposure and
12 mortality.
Confounding by PM and PM Constituents
13 An important question in the evaluation of the association between short-term O3
14 exposure and mortality is whether the relationship is confounded by particulate matter,
15 particularly the PM chemical components that are found in the "summer haze" mixture
16 which also contains O3. However, because of the temporal correlation among these PM
17 components and O3, and their possible interactions, the interpretation of results from
18 multipollutant models that attempt to disentangle the health effects associated with each
19 pollutant is challenging.
20 The potential confounding effects of PM10 and PM2 5 on the O3-mortality relationship
21 were examined by Bell et al. (2007) using data on 98 U.S. urban communities for the
22 years 1987-2000 from the National Morbidity, Mortality, and Air Pollution Study
23 (NMMAPS). In this analysis the authors included PM as a covariate in time-series
24 models, and also examined O3.mortality associations on days when O3 concentrations were
25 below a specified value. This analysis was limited by the small fraction of days when
26 both PM and O3 data were available, due to the every-3rd- or 6th-day sampling schedule
27 for the PM indices, and the limited amount of city-specific data for PM2 5 because it was
28 only collected in most cities since 1999. As a result, of the 91 communities with PM25
29 data, only 9.2% of days in the study period had data for both O3 and PM2 5, resulting in
30 the use of only 62 communities in the PM2 5 analysis. An examination of the correlation
31 between PM (PMi0 and PM2 5) and O3 across various strata of daily PMi0 and PM25
32 concentrations found that neither PM size fraction was highly correlated with O3 across
Draft - Do Not Cite or Quote 6-198 September 2011
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1
2
3
4
5
6
7
9
10
11
12
13
14
15
16
17
any of the strata examined. These results were also observed when using 8-h max and 1-h
max O3 exposure metrics. National and community-specific effect estimates of the
association between short-term O3 exposure and mortality were robust to inclusion of
PMio or PM25 in time-series models through the range of O3 concentrations (i.e.,
<10 ppb, 10-20, 20-40, 40-60, 60-80, and >80 ppb). For example, the percent increases in
nonaccidental deaths per 10 ppb increase 24-h avg O3 concentrations at lag 0-1 day were
0.22% (95% CI: -0.22, 0.65) without PM25 and 0.21% (95% CI: -0.22, 0.64) with PM25
in 62 communities.
Although no strong correlations between PM and O3 were reported by Bell et al. (2007)
the patterns observed suggest regional differences in their correlation. (Table 6-43). Both
PMio and PM2 5 show positive correlations with O3 in the Industrial Midwest, Northeast,
Urban Midwest, and Southeast, especially in the summer months, presumably, because of
the summer peaking sulfate. However, the mostly negative or weak correlations between
PM and O3 in the summer in the Southwest, Northwest, and southern California could be
due to winter-peaking nitrate. Thus, the potential confounding effect of PM on the
O3-mortality relationship could be influenced by the relative contribution of sulfate and
nitrate, which varies regionally and seasonally.
Table 6-43 Correlations between PM and ozone by season and region
No. of Communities
Winter
Spring
Summer
Fall
Yearly
PM10
Industrial Midwest
Northeast
Urban Midwest
Southwest
Northwest
southern California
Southeast
U.S.
19
15
6
9
11
7
25
93
0.37
0.34
0.24
0.00
-0.17
0.19
0.33
0.23
0.44
0.44
0.25
0.02
-0.20
0.08
0.35
0.26
0.44
0.36
0.22
-0.02
-0.13
0.12
0.31
0.24
0.39
0.44
0.26
0.10
-0.11
0.19
0.31
0.26
0.41
0.40
0.24
0.03
-0.16
0.14
0.32
0.25
PM2.5
Industrial Midwest
Northeast
Urban Midwest
Southwest
Northwest
southern California
Southeast
U.S.
19
13
4
9
11
7
26
90
0.18
0.05
0.22
-0.15
-0.32
-0.25
0.38
0.09
0.39
0.26
0.31
-0.08
-0.34
-0.22
0.47
0.21
0.43
0.16
0.15
-0.17
-0.39
-0.25
0.30
0.12
0.44
0.43
0.32
-0.15
-0.24
-0.15
0.37
0.22
0.36
0.25
0.20
-0.14
-0.31
-0.23
0.39
0.16
Source: Bell et al. (2007).
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6-199
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Raw Estimates
15-
10-
o
! 5-
.c
i
1.5-
1.0-
3
I 0.5-
; 0.0 -
-0.5-
-1.0
-0.5
0 5
Without PM10
Posterior Estimates
10
0.0 0.5
Without PM10
1.0
1.5
Source: Reprinted with permission of Informa UK Ltd (Smith et al.. 2009b).
The diagonal line indicates 1:1 ratio.
Figure 6-28 Scatter plots of ozone mortality risk estimates with versus without
adjustment for PMio in NMMAPS cities.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
In an attempt to reassess a number of issues associated with the O3-mortality relationship,
including confounding, Smith et al. (2009b) re-analyzed the publicly available NMMAPS
database for the years 1987-2000. The authors conducted a number of analyses using
constrained distributed lag models and the average of 0- and 1-day lags. In addition,
Smith et al. (2009b) examined the effect of different averaging times (24-h, 8-h, and 1-h
max) on O3-mortality regression coefficients, and whether PM10 confounded the
Os-mortality relationship. The authors reported that, in most cases, O3 mortality risk
estimates were reduced by between 22% and 33% in copollutant models with PM10. This
is further highlighted in Figure 6-28, which shows scatter plots of O3_mortality risk
estimates with adjustment for PM10 versus without adjustment for PM10. Smith et al.
(2009b) point out that a larger fraction (89 out of 93) of the posterior estimates lie below
the diagonal line (i.e., estimates are smaller with PM10 adjustment) compared to the raw
estimates (56 out of 93). This observation could be attributed to both sets of posterior
estimates being calculated by "shrinking towards the mean." However, the most
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1 prominent feature of these plots is that the variation of O3.mortality risk estimates across
2 cities is much larger than the impact of PMi0 adjustment on the O3.mortality relationship.
3 Franklin and Schwartz (2008) examined the sensitivity of O3 mortality risk estimates to
4 the inclusion of PM25 or PM chemical components associated with secondary aerosols
5 (e-g-, sulfate [SO42~], organic carbon [OC], and nitrate [NO3-]) in copollutant models.
6 This analysis consisted of between 3 and 6 years of data from May through September
7 2000-2005 from 18 U.S. communities. The association between O3 and non-accidental
8 mortality was examined in single-pollutant models and after adjustment for PM2 5,
9 sulfate, organic carbon, or nitrate concentrations. The single-city effect estimates were
10 combined into an overall estimate using a random-effects model. In the single-pollutant
11 model, the authors found a 0.89% (95% CI: 0.45, 1.33%) increase in nonaccidental
12 mortality with a 10 ppb increase in same-day 24-h summertime O3 concentrations across
13 the 18 U.S. communities. Adjustment for PM25 mass, which was available for 84% of the
14 days, decreased the O3.mortality risk estimate only slightly (from 0.88% to 0.79%), but the
15 inclusion of sulfate in the model reduced the risk estimate by 31% (from 0.85% to
16 0.58%). However, sulfate data were only available for 18% of the days. Therefore, a
17 limitation of this study is the limited amount of data for PM2 5 chemical components due
18 to the every-3rd-day or every-6th-day sampling schedule. For example, when using a
19 subset of days when organic carbon measurements were available (i.e., 17% of the
20 available days), O3 mortality risk estimates were reduced to 0.51% (95% CI: -0.36 to
21 1.36) in a single-pollutant model.
22 Consistent with the studies previously discussed, the results from Franklin and Schwartz
23 (2008) also demonstrate that the interpretation of the potential confounding effects of
24 copollutants on O3 mortality risk estimates is not straightforward. As presented in Figure
25 6-29, the regional and city-to-city variations in O3 mortality risk estimates appear greater
26 than the impact of adjusting for copollutants. In addition, in some cases, a negative O3
27 mortality risk estimate becomes even more negative with the inclusion of sulfate (e.g.,
28 Seattle) in a copollutant model, or a null O3 mortality risk estimate becomes negative
29 when sulfate is included (e.g., Dallas and Detroit). Thus, the reduction in the overall O3
30 mortality risk estimate (i.e., across cities) needs to be assessed in the context of the
31 heterogeneity in the single-city estimates.
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Seattle
tl Paso
Dallas
Houston
Beaumont
Kansas City
St. Louis
Detroit
Pittsburgh
Buffalo
Rochester
Philadelphia
Boston
1 fa ±H
h-D-^
^C
1 '
rts
1 — ^^
1 ,
D Ozone with sulfate I
x Ozone alone |
1
x n i
r^
j^
=ii
« — i
__i
r-i *. . I
-5 0 5
Percent increase in mortality
with 10 ppb increase in ozone
Source: Reprinted from Franklin and Schwartz (2008).
Figure 6-29 Community-specific ozone-mortality risk estimates for
nonaccidental mortality per 10 ppb increase in same-day 24-h avg
summertime ozone concentrations in single-pollutant models and
copollutant models with sulfate.
1
2
3
4
5
6
7
8
9
10
11
12
13
In the APHENA study, the investigators from the U.S. (NMMAPS), Canadian, and
European (APHEA2) multicity studies collaborated and conducted a joint analysis of
PMio and O3 using each of these datasets (Katsouyanni et al.. 2009). For mortality, each
dataset consisted of a different number of cities and years of air quality data: U.S.
encompassed 90 cities with daily O3 data from 1987-1996 of which 36 cities had summer
only O3 measurements; Europe included 23 cities with 3-7 years of daily O3 data during
1990-1997; and Canada consisted of 12 cities with daily O3 data from 1987 to 1996. As
discussed in Section 6.2.7.2, the APHENA study conducted extensive sensitivity
analyses, of which the 8 df/year results for both the penalized spline (PS) and natural
spline (NS) models are presented in the text for comparison purposes, but only the NS
results are presented in figures because alternative spline models have previously been
shown to result in similar effect estimates (HEI. 2003). Additionally, for the Canadian
results, figures contain risk estimates standardized to both a 40 ppb increment for 1-h
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1 max O3 concentrations, consistent with the rest of the ISA, but also the approximate IQR
2 across the Canadian cities as discussed previously (Section 6.2.7.2).
3 In the three datasets, the authors found generally positive associations between short-term
4 O3 exposure and all-cause, cardiovascular, and respiratory mortality. The estimated
5 excess risks for O3 were larger for the Canadian cities than for the U.S. and European
6 cities. When examining the potential confounding effects of PM10 on O3 mortality risk
7 estimates, the sensitivity of the estimates varied across the data sets and age groups. In
8 the Canadian dataset, adjusting for PM10 modestly reduced O3 risk estimates for all-cause
9 mortality for all ages in the PS (4.5% [95% CI: 2.2, 6.7%]) and NS (4.2% [95% CI: 1.9,
10 6.5%]) models to 3.8% (95% CI: -1.4, 9.8%) and 3.2% (95% CI: -2.2, 9.0%),
11 respectively, at lag 1 for a 40 ppb increase in 1-h max O3 concentrations (Figure 6-30;
12 Table 6-44). However, adjusting for PM10 reduced O3 mortality risk estimates in the >
13 75-year age group, but increased the risk estimates in the <75-year age group. For
14 cardiovascular and respiratory mortality more variable results were observed with O3 risk
15 estimates being reduced and increased, respectively, in copollutant models with PMi0
16 (Figure 6-30; Table 6-44). Unlike the European and U.S. datasets, the Canadian dataset
17 only conducted copollutant analyses at lag 1; as a result, to provide a comparison across
18 study locations only the lag 1 results are presented for the European and U.S. datasets in
19 this section.
20 In the European data, O3 risk estimates were robust when adjusting for PM10 in the year-
21 round data for all-cause, cardiovascular and respiratory mortality. When restricting the
22 analysis to the summer months moderate reductions were observed in O3 risk estimates
23 for all-cause mortality with more pronounced reductions in respiratory mortality. In the
24 U.S. data, adjusting for PM10 moderately reduced O3 risk estimates for all-cause mortality
25 in a year-round analysis at lag 1 (e.g., both the PS and NS models were reduced from
26 0.18%to 0.13%) (Figure 6-30; Table 6-44). Similarto the European data, when
27 restricting the analysis to the summer months, adjusting for PMi0 moderately reduced O3
28 mortality risk estimates in the U.S. However, when examining cause-specific mortality
29 risk estimates, consistent with the results from the Canadian dataset, which employed a
30 similar PM sampling strategy (i.e., every-6th-day sampling), O3 risk estimates for
31 cardiovascular and respiratory mortality were more variable; reduced or increased in
32 all-year and summer analyses. Overall, the estimated O3 risks appeared to be moderately
33 to substantially sensitive to inclusion of PMi0 in copollutant models. Despite the multicity
34 approach, the mostly every-6th-day sampling schedule for PM10 in the Canadian and U.S.
35 datasets greatly reduced the sample size and limits the interpretation of these results.
Draft - Do Not Cite or Quote 6-203 September 2011
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Location
APHENA-U.S.
APHENA-Canada
a
a
a
a
APHENA-Europe
Ages
All
>75
<75
>75
<75
All
All
>75
<75
All
All
>75
<75
>75
<75
All
All-Cause
Cardiovascular
Respiratory
All-Cause
Cardiovascular
Respiratory
-!r-0
All-Year
Summer
All-Year
Summer
All-Year
Summer
All-Year
All-Cause
Cardiovascular
—
_
Respiratory —
—
-10 -5 (
w v
•9- All-Year
-+- Summer
-•— All-Year
3=
^£-
+ Summer
^-0
1 A
^=
!-• All-Year
L-O
^
^ Summer
r-0
3 5 10 15 20 25 30
% Increase
Effect estimates are for a 40 ppb increase in 1 -h max O3 concentrations at lag 1. All estimates are for the 8 df/year model with
natural splines. Circles represent all-year analysis results while diamonds represent summer season analysis results. Open circles
and diamonds represent copollutant models with PM10. Black = all-cause mortality; red = cardiovascular mortality; and blue =
respiratory mortality. An "a" represents risk estimates from APHENA-Canada standardized to an approximate IQR of 5.1 ppb for a
1-h max increase in O3 concentrations (see explanation in Section 6.2.7.2).
Figure 6-30 Percent increase in all-cause (nonaccidental) and cause-specific
mortality from the APHENA study for single- and copollutant
models.
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Table 6-44 Corresponding
Location Mortality
APHENA-U.S. All-Cause
Cardiovascular
Respiratory
APHENA-Canada All-Cause
Cardiovascular
Respiratory
APHENA-Europe All-Cause
Cardiovascular
Respiratory
Effect
Ages
All
>75
<75
>75
<75
All
All
>75
<75
All
All
>75
<75
>75
<75
All
Estimates
Season
All-year
Summer
All-year
Summer
All-year
Summer
All-year
All-year
Summer
All-year
Summer
All-year
Summer
for Figure 6-30
Copollutant
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
PM10
% Increase (95% Cl)
1.42(0.08,2.78)
1.02 (-1.40, 3.50)
4.31 (2.22, 6.45)
1.90 (-0.78, 4.64)
1.10 (-1.33, 3.67)
0.47 (-4.61, 5.79)
-0.1 6 (-3.02, 2.86)
1.34 (-3.63, 6.61)
3.58 (0.87, 6.37)
-1.1 7 (-6.18, 4.07)
3.18(0.31,6.12)
1 .26 (-4.46, 7.28)
2.46 (-1.87, 6.86)
3.50 (-4.23, 11.8)
6.04(1.18,11.1)
7.03 (-3.48, 18.5)
4.15(1.90,6.45)
0.52 (0.24, 0.80)a
3.18 (-2.18, 8.96)
0.40 (-0.28, 1.1 0)a
5.62 (0.95, 10.7)
0.70(0.12,1.30)3
1.90 (-9.03, 14.1)
0.24 (-1.20,1.70)3
1.10 (-4.08, 6.61)
0.1 4 (-0.53, 0.82)3
-2.64 (-14.7, 11.5)
-0.34 (-2.00, 1.40)3
0.87 (-6.40, 8.96)
0.11 (-0.84,1.10)3
22.3 (-12.6, 71.3)
2.60 (-1.70, 7.10)3
1.02(0.39,1.66)
1.26(0.47,1.98)
2.06(1.10,2.94)
1.26(0.16,2.30)
1.10 (-0.47, 2.70)
1.1 8 (-0.55, 2.94)
1 .34 (-0.24, 2.94)
1.74 (-0.31, 3.75)
2.54 (0.39, 4.80)
1.58 (-0.70, 3.99)
1.66 (-0.70, 4.15)
1.66 (-1.02, 4.40)
1.42 (-1.02, 3.83)
1.42 (-1.02, 3.83)
4.31 (1.66,7.11)
1.18 (-1.79, 4.31)
1
2
"Risk estimates from APHENA-Canada standardized to an approximate IQR of 5.1 ppb for a 1-h max increase in O3
concentrations (see explanation in Section 6.2.7.2).
Stafoggia et al. (2010) examined the potential confounding effects of PMi0 on the
O3-mortality relationship in individuals 35 years of age and older in 10 Italian cities from
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1 2001 to 2005. In a time-stratified case-crossover analysis, using data for the summer
2 months (i.e., April-September), the authors examined O3-mortality associations across
3 each city, and then obtained a pooled estimate through a random-effects meta-analysis.
4 Stafoggia et al. (2010) found a strong association with nonaccidental mortality (9.2%
5 [95% CI: 5.4, 13.0%] for a 30 ppb increase in 8-h max O3 concentrations) in an
6 unconstrained distributed lag model (lag 0-5) that persisted in copollutant models with
7 PM10 (9.2% [95% CI: 5.4, 13.7%]). Additionally, when examining cause-specific
8 mortality, the authors found positive associations between short-term O3 exposure and
9 cardiovascular (14.3% [95% CI: 6.7, 22.4%]), cerebrovascular (8.5% [95% CI: 0.1,
10 16.3%]), and respiratory (17.6% [95% CI: 1.8, 35.6%]) mortality in single-pollutant
11 models. In copollutant models, O3-mortality effect estimates for cardiovascular and
12 cerebrovascular mortality were robust to the inclusion of PMi0 (9.2% [95% CI: 5.4,
13 13.7%]) and 7.3% [95% CI: -1.2, 16.3%], respectively), and attenuated, but remained
14 positive, for respiratory mortality (9.2% [95% CI: -6.9, 28.8%]). Of note, the correlations
15 between O3 and PM10 across cities were found to be generally low, ranging from (-0.03 to
16 0.49). The authors do not specify the sampling strategy used for PMi0 in this analysis.
Confounding by Seasonal Trend
17 The APHENA study (Katsouyanni et al.. 2009). mentioned above, also conducted
18 extensive sensitivity analyses to identify the appropriate: smoothing method and basis
19 functions to estimate smooth functions of time in city-specific models; and degrees of
20 freedom to be used in smooth functions of time, to adjust for seasonal trends. Because O3
21 peaks in the summer and mortality peaks in the winter, not adjusting or not sufficiently
22 adjusting for the seasonal trend would result in an apparent negative association between
23 the O3 and mortality time-series. Katsouyanni et al. (2009) examined the effect of the
24 extent of smoothing for seasonal trends by using models with 3 df/year, 8 df/year (the
25 choice for their main model), 12 df/year, and df/year selected using the sum of absolute
26 values of partial autocorrelation function of the model residuals (PACF) (i.e., choosing
27 the degrees of freedom that minimizes positive and negative autocorrelations in the
28 residuals). Table 6-45 presents the results of the degrees of freedom analysis using
29 alternative methods to calculate a combined estimate: the Berkey et al. (1998) meta-
30 regression and the two-level normal independent sampling estimation (TLNISE)
31 hierarchical method. The results show that the methods used to combine single-city
32 estimates did not influence the overall results, and that neither 3 df/year nor choosing the
3 3 df/year by minimizing the sum of absolute values of PACF of regression residuals was
34 sufficient to adjust for the seasonal negative relationship between O3 and mortality.
35 However, it should be noted, the majority of studies in the literature that examined the
36 mortality effects of short-term O3 exposure, particularly the multicity studies, used 7 or
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1 8 df/year to adjust for seasonal trends, and in both methods a positive association was
2 observed between O3 exposure and mortality.
Table 6-45 Sensitivity of ozone risk estimates per 10 |jg/m3 increase in
24-h avg ozone concentrations at lag 0-1 to alternative methods for
adjustment of seasonal trend, for all-cause mortality using Berkey
MLE and TLNSE Hierarchical Models
Seasonality Control
3 df/year
8 df/year
12 df/year
PACF
Berkey
-0.54 (-0.88, 0.20)
0.30(0.11,0.50)
0.34(0.15,0.53)
-0.62 (-1 .01 , -0.22)
TLNISE
-0.55 (-0.88, -0.22)
0.31 (0.09, 0.52)
0.33(0.12,0.54)
-0.62 (-0.98, -0.27)
Source: Reprinted with permission of Health Effects Institute (2009).
6.6.2.2 Effect Modification
3 There have been several multicity studies that examined potential effect modifiers, or
4 time-in variant factors, that may modify O3 mortality risk estimates. These effect
5 modifiers can be categorized into either individual-level or community-level
6 characteristics, which are traditionally, examined in second stage regression models. The
7 results from these analyses also inform upon whether certain populations are susceptible
8 to O3-related health effects (Chapter 8). In addition to potentially modifying the
9 association between short-term O3 exposure and mortality, both individual-level and
10 community-level characteristics may also contribute to the apparent geographic pattern of
11 spatial heterogeneity in O3 mortality risk estimates. As a result, the geographic pattern of
12 O3 mortality risk estimates is also evaluated in this section.
Individual-Level Characteristics
13 Medina-Ramon and Schwartz (2008) conducted a case-only study in 48 U.S. cities to
14 identify populations potentially susceptible to O3-related mortality for the period
15 1989-2000 (May through September of each year [i.e., warm season]). A case-only
16 design predicts the occurrence of time-invariant characteristics among cases as a function
17 of the exposure level (Armstrong. 2003). For each potential effect modifier (time-
18 invariant individual-level characteristics), city-specific logistic regression models were
19 fitted, and the estimates were pooled across all cities. Furthermore, the authors examined
20 potential differences in individual effect modifiers according to several city
21 characteristics (e.g., mean O3 level, mean temperature, households with central air
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1 conditioning, and population density) in a meta-regression. Across cities the authors
2 found a 1.96% (95% CI: 1.14-2.82%) increase in mortality at lag 0-2 for a 30 ppb
3 increase in 8-h max O3 concentrations. Additionally, Medina-Ramon and Schwartz
4 (2008) examined a number of individual-level characteristics (e.g., age, race) and chronic
5 conditions (e.g., secondary causes of death) as effect modifiers of the association between
6 short-term O3 exposure and mortality. The authors found that older adults (i.e., > 65),
7 women >60 years of age, black race, and secondary atrial fibrillation showed the greatest
8 additional percent change in O3-related mortality (Table 6-46). In addition, when
9 examining city-level characteristics, the authors found that older adults, black race, and
10 secondary atrial fibrillation had a larger effect on O3 mortality risk estimates in cities with
11 lower O3 levels. Of note, a similar case-only study (Schwartz. 2005b) examined potential
12 effect modifiers of the association between temperature and mortality, which would be
13 expected to find results consistent with the Medina-Ramon and Schwartz (2008) study
14 due to the high correlation between temperature and O3. However, when stratifying days
15 by temperature Schwartz (2005b) found strong evidence that diabetes modified the
16 temperature-mortality association on hot days, which was not as evident when examining
17 the O3-mortality association in Medina-Ramon and Schwartz (2008). This difference
18 could be due to the study design and populations included in both studies, a multeity
19 study including all ages (Medina-Ramon and Schwartz. 2008) compared to a single-city
20 study of individuals > 65 years of age (Schwartz. 2005b). However, when examining
21 results stratified by race, nonwhites were found to have higher mortality risks on both hot
22 and cold days, which provide some support for the additional risk found for black race in
23 Medina-Ramon and Schwartz (2008).
24 Individual-level factors that may result in susceptibility to O3-related mortality were also
25 examined by Stafoggia et al. (2010). As discussed above, using a time-stratified case-
26 crossover analysis, the authors found an association between short-term O3 exposure and
27 nonaccidental mortality in an unconstrained distributed lag model in 10 Italian cities
28 (9.2% [95% CI: 5.4, 13.0%; lag 0-5 for a 30 ppb increase in 8-h max O3 concentrations).
29 Stafoggia et al. (2010) conducted additional analyses to examine whether age, sex,
30 income level, location of death, and underlying chronic conditions increased the risk of
31 O3-related mortality, but data were only available for nine of the cities for these analyses.
32 Of the individual-level factors examined, the authors found the strongest evidence for
33 increased risk of O3-related mortality in individuals > 85 years of age (22.4% [95% CI:
34 15.0, 30.2%]), women (13.7% [95% CI: 8.5, 19.7%]), and out-of-hospital deaths (13.0%
35 [95% CI: 6.0, 20.4%]). When focusing specifically on out-of hospital deaths and the
36 subset of individuals with chronic conditions, Stafoggia et al. (2010) found the strongest
37 association for individuals with diabetes, which is consistent with the potentially
38 increased susceptibility of diabetics on hot days observed in Schwartz (2005b).
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Table 6-46 Additional percent change in ozone-related mortality for individual-
level susceptibility factors
Percentage
(95% Cl)
Socio-demographic characteristics
Age 65 yr or older
Women
Women <60 yr old"
Women > 60 yr old"
Black race
Low education
1.10
0.58
-0.09
0.60
0.53
-0.29
0.44,1.77
0.18,0.98
-0.76, 0.58
0.25, 0.96
0.19,0.87
-0.81,0.23
Chronic conditions (listed as secondary cause)
Respiratory system diseases
Asthma
COPD
1.35
0.01
-0.31,3.03
-0.49, 0.52
Circulatory system diseases
Atherosclerosis
Atherosclerotic CVD
Atherosclerotic heart disease
Congestive heart disease
Atrial fibrillation
Stroke
-0.72
0.74
-0.38
-0.04
1.66
0.17
-1.89,0.45
-0.86, 2.37
-1.70,0.96
-0.39, 0.30
0.03, 3.32
-0.28, 0.62
Other diseases
Diabetes
Inflammatory diseases
0.19
0.18
-0.46, 0.84
-1 .09, 1 .46
'These estimates represent the additional percent change in mortality for persons who had the characteristic being examined compared to
persons who did not have the characteristic, when the mean 03 level of the previous 3 days increased 10 ppb. These values were not standardized
because they do not represent the actual effect estimate for the characteristic being evaluated, but instead, the difference between effect estimates
for persons with versus without the condition.
""Compared with males in the same age group.
Source: Reprinted with permission from Lippincott Williams & Wilkins, Medina-Ramon and Schwartz (2008).
1 Additionally, Cakmak et al. (2011) examined the effect of individual-level characteristics
2 that may modify the O3-mortality relationship in 7 Chilean cities. In a time-series analysis
3 using a constrained distributed lag of 0-6 days, Cakmak et al. (2011) found evidence for
4 larger O3 mortality effects in individuals > 75 years of age compared to younger ages,
5 which is similar to Medina-Ramon and Schwartz (2008) and Stafoggia et al. (2010).
6 Unlike the studies discussed above O3-mortality risk estimates were found to be slightly
7 larger in males (3.71% [95% CI: 0.79, 6.66] for a 40 ppb increase in max 8-h avg O3
8 concentrations), but were not significantly different than those observed for females
9 (3.00% [95% CI: 0.43, 5.68]). The major focus of Cakmak et al. (2011) is the
10 examination of the influence of SES indicators (i.e., educational attainment, income level,
11 and employment status) on the O3-mortality relationship. The authors found the largest
12 risk estimates in the lowest SES categories for each of the indicators examined this
13 includes: primary school not completed when examining educational attainment; the
14 lowest quartile of income level; and unemployed individuals when comparing
15 employment status.
16 Overall, uncertainties exist in the interpretation of the potential effect modifiers,
17 identified in Medina-Ramon and Schwartz (2008). Stafoggia et al. (2010). and Cakmak et
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1 al. (2011) of the O3-mortality relationship due to the expected heterogeneity in O3 mortal-
2 ity risk estimates across cities as highlighted in Smith et al. (2009b) (Figure 6-28) and
3 Franklin and Schwartz (2008) (Figure 6-29). For example, it is difficult to determine the
4 relative importance of a susceptibility factor that results in an additional percent increase
5 in mortality in a multicity analysis when analyses of the individual cities within the study
6 did not indicate associations between O3 and mortality. In addition, it is likely that
7 individual-level susceptibility factors identified in Medina-Ramon and Schwartz (2008).
8 Stafoggia et al. (2010). and Cakmak et al. (2011) only modify the O3-mortality relation-
9 ship. The factors identified span pollutants as is evident by older adults (i.e., > 65) often
10 being identified as an effect modifier of PM mortality risk estimates (U.S. EPA. 2009d).
Community-level Characteristics
11 Several studies also examined city-level (i.e., ecological) variables to explain city-to-city
12 variation in estimated O3 mortality risk estimates. Bell and Dominici (2008) investigated
13 whether community-level characteristics, such as race, income, education, urbanization,
14 transportation use, PM and O3 levels, number of O3 monitors, weather, and air
15 conditioning use could explain the heterogeneity in O3-mortality risk estimates across
16 cities. The authors analyzed 98 U.S. urban communities from NMMAPS for the period
17 1987-2000. In the all-year regression model that included no community-level variables,
18 a 20 ppb increase in 24-h avg O3 concentrations during the previous week was associated
19 with a 1.04% (95% CI: 0.56, 1.55) increase in mortality. Bell and Dominci (2008) found
20 that higher O3.mortality effect estimates were associated with higher: percent
21 unemployment, fraction of the population Black/African-American, percent of the
22 population that take public transportation to work; and with lower: temperatures and
23 percent of households with central air conditioning (Figure 6-31). The modification of
24 O3-mortality risk estimates reported for city-specific temperature and prevalence of
25 central air conditioning in this analysis confirm the result from the meta-analyses
26 reviewed in the 2006 O3 AQCD.
27 The APHENA project (Katsouvanni et al.. 2009) examined potential effect modification
28 of O3 risk estimates in the Canadian, European, and U.S. data sets using a consistent set
29 of city-specific variables. Table 6-47 presents the results from all age analyses for all-
30 cause mortality using all-year O3 data for the average of lag 0-1 day. While there are
31 several significant effect modifiers in the U.S. data, the results are mostly inconsistent
32 with the results from the Canadian and European data sets. The positive effect
33 modification by percentage unemployed and the negative effect modification by mean
34 temperature (i.e., a surrogate for air conditioning rate) are consistent with the results
35 reported by Bell and Dominici (2008) discussed above. However, the lack of consistency
36 across the data sets, even between the Canadian and U.S. data, makes it difficult to
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1
2
3
interpret the results. Some of these associations may be due to coincidental correlations
with other unmeasured factors that vary regionally (e.g., mean SO2 tend to be higher in
the eastern U.S.).
fl 4-
I g
i§
*.l 2-
fi
I
fl
|s.2J
I* «i
:: "
i§
' *.->, 00° O 0 fc 2-
• sf-£> jQ £2)_£_— -® — *" c S
_^L-22&a~~£>/^rf~°'° °» "6 !
u. " i j i"^ o Co^f a ti y n
O° ?t»o«,to^ &-E
, o^ ' « S -o
°O O oj 0-
o* ° iSjJ
0
o
6 »
* « o» O ° n
o o e 1° 0 W0 O °___— — —
t> lO^xO— °" — ^^ —
"^^>i*rfc * ° ° °
fs^t*T!t> *X«« ».
•7*fv& ^g * o
•#
CO
00 °°
"1 .• °-| .
3 4 5 6 7 B
Percentage of population unemployed
10 20 30 40 50
Percentage of population
Black/African American
tl *
;.e 2 •
0 -
e -2
50 55 60 65 70 75
Long-term temperature (°F)
10 20 30 40 50
Percentage of population taking
public transportaton to work
r§ 4
i
is
S 2-
It
0 20 40 60 80
Percentage of households with central AC
Source: Reprinted with permission of Johns Hopkins Bloomberg School of Public Health, Bell and Dominici (2008).
Figure 6-31 Ozone mortality risk estimates and community-specific
characteristics, U.S., 1987-2000. The size of each circle
corresponds to the inverse of the standard error of the
community's maximum likelihood estimate. Risk estimates are for a
10 ppb increase in 24-h avg ozone concentrations during the
previous week.
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Table 6-47 Percent change in all-cause mortality, for all ages, associated with
a 40ppb increase in 1-h max ozone concentrations at Lag 0-1 at the
25th and 75th percentile of the center-specific distribution of
selected effect modifiers
Canada
Effect
Modifier
N02CV
Mean S02
03CV
Mean
N02/PM10
Mean
Temperature
% > 75 yr
Age
standardized
Mortality
%
Unemployed
25th Percentile
Estimate
(95% Cl)
3.10
(1 .90, 4.40)
2.22
(0.71 , 3.83)
2.86
(0.79, 5.05)
3.91
(2.54, 5.29)
2.86
(0.95, 4.72)
2.22
(0.79, 3.58)
2.62
(0.79, 4.48)
2.78
(1.42,4.07)
75th
Percentile t
Estimate Value
(95% Cl)
3.99 1 .33
(2.38, 5.62)
4.72 2.16
(2.94, 6.61)
3.50 0.60
(2.14,4.89)
2.54 -1.58
(0.95,4.15)
3.50 0.83
(2.22, 4.89)
4.23 2.68
(3.02, 5.54)
4.07 1.14
(2.22, 5.87)
3.75 1 .88
(2.54, 4.89)
Europe
25th Percentile
Estimate
(95% Cl)
1.66
(0.71,2.62)
1.58
(0.47, 2.62)
2.62
(1 .50, 3.75)
1.74
(0.87, 2.70)
1.58
(0.39, 2.86)
1.50
(0.55, 2.46)
1.10
(-0.16,2.38)
1.42
(-0.47, 3.34)
75th
Percentile t
Estimate Value
(95% Cl)
1 .34 -0.49
(-0.08,
2.78)
1.66 0.16
(0.39, 2.86)
1.10 -2.65
(0.24, 1 .98)
1 .50 -0.43
(0.47, 2.62)
1 .58 -0.04
(0.31,2.78)
1 .82 0.52
(0.55,3.10)
1 .98 1 .07
(0.79, 3.26)
1 .34 -0.07
(-0.47,
3.18)
U
25th Percentile
Estimate
(95% Cl)
1.26
(0.47, 1 .98)
0.47
(-0.47, 1 .42)
0.16
(-0.70,1.10)
-0.08
(-1 .02, 0.95)
2.14
(1.34,2.94)
1.02
(0.24, 1 .90)
0.00
(-0.94, 0.87)
0.16
(-0.78,1.18)
.S.
75th
Percentile
Estimate
(95% Cl)
0.08
(-0.78,
0.95)
1.98
(1.10,2.94)
1.50
(0.71,2.22)
1.26
(0.47, 2.06)
0.00
(-0.78,
0.79)
1.02
(0.31,1.74)
1.58
(0.87, 2.38)
1.50
(0.71,2.30)
t
Value
-2.87
2.79
2.68
2.64
-4.40
-0.02
3.81
2.45
Source: Adapted with permission of Health Effects Institute, Katsouyanni et al. (2009).
1
2
3
4
5
6
7
8
9
10
Regional Pattern of Ozone-Mortality Risk Estimates
In addition to examining whether individual- and community-level factors modify the
O3.mortality association, studies also examined whether these associations varied
regionally within the U.S. Bell and Dominici (2008). in the study discussed above, also
noted that O3-mortality risk estimates were higher in the Northeast (1.44% [95% Cl: 0.78,
2.10%]) and Industrial Midwest (0.73% [95% Cl: 0.11, 1.35%]), while null associations
were observed in the Southwest and Urban Midwest (Table 6-48). The regional
heterogeneity in O3-mortality risk estimates was further reflected by Bell and Dominici
(2008) in a map of community-specific Bayesian O3-mortality risk estimates (Figure 6-
32). It is worth noting that in the analysis of PMi0 using the same data set, Peng et al.
(2005) also found that both the Northeast and Industrial Midwest showed particularly
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
elevated effects, especially during the summer months. As mentioned above, although no
evidence for confounding of O3 mortality risk estimates by PMi0 was observed, Bell et al.
(2007) did find regional differences in the correlation between O3 and PM10. Thus, the
heterogeneity in O3 mortality risk estimates may need to be examined as a function of the
correlation between PM and O3.
Smith et al. (2009b), as discussed earlier, also examined the regional difference in O3
mortality risk estimates across the same seven regions and similarly found evidence for
regional heterogeneity. In addition, Smith et al. (2009b) constructed spatial maps of the
risk estimates by an extension of a hierarchical model that allows for spatial auto-
correlation among the city-specific random effects. Figure 6-31 presents the spatial map
of O3 mortality coefficients from the Smith et al. (2009b) analysis that used 8-h max O3
concentrations during the summer. The results from the Bell and Dominici (2008)
analysis (Figure 6-32) shows much stronger apparent heterogeneity in O3-mortality risk
estimates across cities than the smoothed map from Smith et al. (2009b) (Figure 6-33),
but both maps generally show larger risk estimates in the eastern region of the U.S.
Table 6-48 Percentage increase in daily mortality for a 10 ppb increase in 24-h
avg ozone concentrations during the previous week by geographic
region in the U.S., 1987-2000
No. of Communities
Regional Estimate
95% PI*
Regional results
Industrial Midwest
Northeast
Northwest
southern California
Southeast
Southwest
Urban Midwest
20
16
12
7
26
9
7
0.73
1.44
0.08
0.21
0.38
-0.06
-0.05
0.11,1.35
0.78,2.10
-0.92, 1 .09
-0.46, 0.88
-0.07, 0.85
-0.92, 0.81
-1.28,1.19
National results
All continental communities
All communities
97
98
0.51
0.52
0.27, 076
0.28, 0.77
Source: Used with permission from Johns Hopkins Bloomberg School of Public Health, Bell and Dominici (2008).
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Source: Reprinted with permission of Johns Hopkins Bloomberg School of Public Health (Bell and Dominici, 2008).
Figure 6-32 Community-specific Bayesian ozone-mortality risk estimates in 98
U.S. communities.
8H: summer
- 1.0
0.8
- 0.6
0.4
-0.2
-0.0
Source: Reprinted with permission of Informa UK Ltd. (Smith et al.. 2009b).
Figure 6-33 Map of spatially dependent ozone-mortality coefficients for 8-h max
ozone concentrations using summer data.
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6.6.2.3 Interaction
1 The terms effect modification and interaction are often used interchangeably, but
2 theoretically they represent different concepts. Although interactions can lead to either
3 antagonistic or synergistic effects, most studies attempt to identify potential factors that
4 interact synergistically with O3 to increase the risk of mortality. Within this section,
5 interactive effects are defined as time-varying covariates, such as temperature and
6 copollutants that are included in 1st stage time-series regression models. To date, only a
7 few time-series studies have investigated the potential interaction between O3 exposure
8 and copollutants or weather variables. This can be attributed to the moderate to high
9 correlation between O3 and these covariates, which makes such investigations
10 methodologically challenging.
11 Ren et al. (2008) examined the possible synergistic effect between O3 and temperature on
12 mortality in the 60 largest eastern U.S. communities from the NMMAPS data during the
13 warm months (i.e., April to October) from 1987-2000. This analysis was restricted to the
14 eastern areas of the U.S. (i.e., Northeast, Industrial Midwest and Southeast) because a
15 previous study which focused specifically on the eastern U.S. found that
16 temperature-mortality patterns differ between the northeast and southeast regions
17 possibly due to climatic differences (Curriero et al.. 2002). To examine possible
18 geographic differences in the interaction between temperature and O3, Ren et al. (2008)
19 further divided the NMMAPS regions into the Northeast, which included the Northeast
20 and Industrial Midwest regions (34 cities), and the Southeast, which included the
21 Southeast region (26 cities). The potential synergistic effects between O3 and temperature
22 were examined using two different models. Model 1 included an interaction term in a
23 Generalized Additive Model (GAM) for O3 and maximum temperature (3-day avg values
24 were used for both terms) to examine the bivariate response surface and the pattern of
25 interaction between the two variables in each community. Model 2 consisted of a
26 Generalized Linear Model (GLM) that used interaction terms to stratify by "low,"
27 "moderate," and "high" temperature days using the first and third quartiles of temperature
28 as cut-offs to examine the percent increase in mortality in each community. Furthermore,
29 a two-stage Bayesian hierarchical model was used to estimate the overall percent increase
30 in all-cause mortality associated with short-term O3 exposure across temperature levels
31 and each region using model 2. The same covariates were used in both model 1 and 2.
32 The bivariate response surfaces from model 1 suggest possible interactive effects
33 between O3 and temperature although the interpretation of these results is not
34 straightforward due to the high correlation between these terms. The apparent interaction
35 between temperature and O3 as evaluated in model 2 varied across geographic regions. In
36 the northeast region, a 20 ppb increase in 24-h avg O3 concentrations at lag 0-2 was
37 associated with an increase of 4.49% (95% posterior interval [PI]: 2.39, 6.36%), 6.21%
Draft - Do Not Cite or Quote 6-215 September 2011
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1 (95% PI: 4.47, 7.66%) and 12.8% (95% PI: 9.77, 15.7%) in mortality at low, moderate
2 and high temperature levels, respectively. The corresponding percent increases in
3 mortality in the southeast region were 2.27% (95% PI: -2.23, 6.46%) for low temperature,
4 3.02% (95% PI: 0.44, 5.70%) for moderate temperature, and 2.60% (95% PI: -0.66,
5 6.01%) for high temperature.
6 When examining the relationship between temperature and O3-related mortality, the
7 results reported by Ren et al. (2008) (i.e., higher O3-mortality risks on days with higher
8 temperatures) may appear to contradict the results of Bell and Dominici (2008) described
9 earlier (i.e., communities with higher temperature have lower O3-mortality risk
10 estimates). However, the observed difference in results can be attributed to the
11 interpretation of effect modification in a second-stage regression which uses long-term
12 average temperatures, as was performed by Bell and Dominici (2008). compared to a
13 first-stage regression that examines the interaction between daily temperature and O3-
14 related mortality. In this case, the second-stage regression results from Bell and Dominici
15 (2008) indicate that a city with lower temperatures, on average, tend to show a stronger
16 O3 mortality effect, whereas, in the first-stage regression performed by Ren et al. (2008).
17 the days with higher temperature tend to show a larger O3-mortality effect. This observed
18 difference may in part reflect the higher air conditioning use in communities with higher
19 long-term average temperatures. Therefore, the findings from Ren et al. (2008) indicating
20 generally lower O3 risk estimates in the southeast region where the average temperature is
21 higher than in the northeast region is consistent with the regional results reported by Bell
22 and Dominici (2008). As demonstrated by the results from both Ren et al. (2008) and
23 Bell and Dominici (2008) caution is required when interpreting results from studies that
24 examined interactive effects using two different approaches because potential effect
25 modification as suggested in a second-stage regression generally does not provide
26 evidence for a short-term interaction examined in a first-stage regression. Overall, further
27 examination of the potential interactive (synergistic) effects of O3 and covariates in time-
28 series regression models is required to more clearly understand the factors that may
29 influence O3 mortality risk estimates.
6.6.2.4 Evaluation of the Ozone-Mortality C-R Relationship and
Related Issues
30 Evaluation of the O3-mortality concentration-response relationship is not straightforward
31 because the evidence from multicity studies (using log-linear models) suggests that
32 O3-mortality associations are highly heterogeneous across regions. In addition, there are
33 numerous issues that may influence the shape of the O3-mortality concentration-response
34 relationship that warrant examination including: multi-day effects (distributed lags),
Draft - Do Not Cite or Quote 6-216 September 2011
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1 potential adaptation, mortality displacement (i.e., hastening of death by a short period),
2 and the exposure metric used to compute risks (e.g., 1-h daily max versus 24-h avg). The
3 following section presents the recent studies identified that conducted an initial
4 examination of these issues.
Multiday Effects, Mortality Displacement, and Adaptation
5 The pattern of positive lagged associations followed by negative associations in a
6 distributed lag model may be considered an indication of "mortality displacement" (i.e.,
7 deaths are occurring in frail individuals and exposure is only moving the day of death to a
8 day slightly earlier). Zanobetti and Schwartz (2008b) examined this issue in 48 U.S. cities
9 during the warm season (i.e., June-August) for the years 1989-2000. In an initial analysis,
10 the authors applied a GLM to examine same-day O3-mortality effects, and in the model
11 included an unconstrained distributed lag for apparent temperature to take into account
12 the effect of temperature on the day death occurred and the previous 7 days. To examine
13 mortality displacement Zanobetti and Schwartz (2008b) refit models using two
14 approaches: an unconstrained and a smooth distributed lag each with 21-day lags for O3.
15 In this study, all-cause mortality as well as cause-specific mortality (i.e., cardiovascular,
16 respiratory, and stroke) were examined for evidence of mortality displacement. The
17 authors found a 0.96% (95% CI: 0.60, 1.30%) increase in all-cause mortality across all 48
18 cities for a 30 ppb increase in 8-h max O3 concentrations at lag 0 whereas the combined
19 estimate of the unconstrained distributed lag model (lag 0-20) was 1.54% (95% CI: 0.15,
20 2.91%). Similarly, when examining the cause-specific mortality results (Table 6-49),
21 larger risk estimates were observed for the distributed lag model compared to the lag
22 0 day estimates. However, for stroke a slightly larger effect was observed at lags 4-20
23 compared to lags 0-3 suggesting a larger window for O3-induced stroke mortality. This is
24 further supported by the sum of lags 0 through 20 days showing the greatest effect.
25 Overall, these results suggest that estimating the mortality risk using a single day of O3
26 exposure may underestimate the public health impact, but the extent of multi-day effects
27 appear to be limited to a few days. This is further supported by the shape of the combined
28 smooth distributed lag (Figure 6-34). It should be noted that the proportion of total
29 variation in the effect estimates due to the between-cities heterogeneity, as measured by
30 I2 statistic, was relatively low (4% for the lag 0 estimates and 21% for the distributed
31 lag), but 21 out of the 48 cities exhibited null or negative estimates. As a result, the
32 estimated shape of the distributed lag cannot be interpreted as a general form of lag
33 structure of associations applicable to all the cities included in this analysis.
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Table 6-49 Estimated effect of a 10 ppb increase in 8-h max
concentrations on mortality during the summer
day and distributed lag models
% (Percentage)
ozone
months for single-
95% Cl
Total mortality
LagO
Sum lags 0-20
Sum lags 0-3
Sum lags 4-20
0.32
0.51
0.53
-0.02
0.20, 0.43
0.05, 0.96
0.28, 0.77
-0.35, 0.31
Cardiovascular mortality
LagO
Sum lags 0-20
Sum lags 0-3
Sum lags 4-20
0.47
0.49
0.80
-0.23
0.30, 0.64
-0.01 , 1 .00
0.48,1.13
-0.67, 0.22
Respiratory mortality
LagO
Sum lags 0-20
Sum lags 0-3
Sum lags 4-20
0.54
0.61
0.83
-0.24
0.26, 0.81
-0.41 , 1 .65
0.38,1.28
-1.08,0.60
Stroke
LagO
Sum lags 0-20
Sum lags 0-3
Sum lags 4-20
0.37
2.20
0.92
1.26
0.01,0.74
0.76, 3.67
0.26,1.59
0.05, 2.49
Source: Reprinted with permission from American Thoracic Society, Zanobetti and Schwartz (2008b).
1 Samoli et al. (2009) also investigated the temporal pattern of mortality
2 effects in response to short-term exposure to O3 in 21 European cities that were included
3 in the APHEA2 project. Using a method similar to Zanobetti and Schwartz (2008b), the
4 authors applied unconstrained distributed lag models with lags up to 21 days in each city
5 during the summer months (i.e., June through August) to examine the effect of O3 on all-
6 cause, cardiovascular, and respiratory mortality. They also applied a generalized additive
7 distributed lag model to obtain smoothed distributed lag coefficients. However, unlike
8 Zanobetti and Schwartz (2008b). Samoli et al. (2009) controlled for temperature using a
9 linear term for humidity and an unconstrained distributed lag model of temperature at
10 lags 0-3 days. The choice of 0- through 3-day lags of temperature was based on a
11 previous European multicity study (Baccini et al.. 2008). which suggested that summer
12 temperature effects last only a few days. Upon combining the individual city estimates
13 across cities in a second stage regression, Samoli et al. (2009) found that the estimated
14 effects on respiratory mortality were extended for a period of two weeks. However, for
15 all-cause and cardiovascular mortality, the 21-day distributed lag models yielded null or
16 (non-significant) negative estimates (Table 6-50). Figure 6-35 shows the distributed lag
17 coefficients for all-cause mortality, which exhibit a declining trend and negative
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1
2
3
4
5
coefficients beyond 5-day lags. The authors' interpretation of these results was that
"using single-day exposures may have overestimated the effects on all-cause and
cardiovascular mortality, but underestimated the effects on respiratory mortality." Thus,
the results in part suggest evidence of mortality displacement for all-cause and
cardiovascular mortality.
<"*
o'
0)
0>
b
o
o
CN|
d '
10
15
Day Lag
20
Source: Reprinted with permission of American Thoracic Society (Zanobetti and Schwartz. 2008b).
The triangles represent the percent increase in all-cause mortality for a 10 ppb increase in 8-h max ozone concentrations at each
lag while the shaded areas are the 95% point-wise confidence intervals.
Figure 6-34 Estimated combined smooth distributed lag for 48 U.S. cities
during the summer months.
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September 2011
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Table 6-50 Estimated percent increase in cause-specific mortality (and 95%
CIs) for a 10-ug/m3 increase in maximum 8-h ozone during
June-August, for the same day (lag 0), the average of the same and
previous day (lag 0-1), the unconstrained distributed lag model for
the sum of 0-20 days and the penalized distributed lag model
(lag 0-20)
Fixed effects
% (95% Cl)
Random effects
% (95% Cl)
Total mortality
LagO
Average lags 0-1
Sum lags 0-20, unconstrained
Sum lags 0-20, penalized
0.28(0.11,0.45)
0.24(0.15,0.34)
0.01 (-0.40, 0.41)
0.01 (-0.41 , 0.42)
0.28 (0.07, 0.48)
0.22 (0.08, 0.35)
-0.54 (-1.28, 0.20)
-0.56 (-1.30, 0.1 9)
Cardiovascular mortality
LagO
Average lags 0-1
Sum lags 0-20, unconstrained
Sum lags 0-20, penalized
0.43(0.18,0.69)
0.33(0.19,0.48)
-0.33 (-0.93, 0.29)
-0.32 (-0.92, 0.28)
0.37 (0.05, 0.69)
0.25 (0.03, 0.47)
-0.62 (-1.47, 0.24)
-0.57 (-1.39, 0.26)
Respiratory mortality
LagO
Average lags 0-1
Sum lags 0-20, unconstrained
Sum lags 0-20, penalized
0.36 (-0.21, 0.94)
0.40(0.11,0.70)
3.35(1.90,4.83)
3.66 (2.25, 5.08)
0.36 (-0.21 , 0.94)
0.40(0.11,0.70)
3.35(1.90,4.83)
3.66 (2.25, 5.08)
Source: Used with permission from BMJ Group (Samoli etal., 2009).
1 Although the APHENA project (Katsouyanni et al.. 2009) did not specifically investigate
2 mortality displacement and therefore did not consider longer lags (e.g., lag > 3 days), the
3 study did present O3 risk estimates for lag 0-1, lag 1, and a distributed lag model of 0-
4 2 days in the Canadian, European, and U.S. datasets. Katsouyanni et al. (2009) found that
5 the results vary somewhat across the regions, but, in general, there was no indication that
6 the distributed lag model with up to a 2-day lag yielded meaningfully larger O3 mortality
7 risk estimates than the lag 0-1 and lag 1 results. For example, for all-cause mortality,
8 using the model with natural splines and 8 df/year to adjust for seasonal trends, a reported
9 percent excess risk for mortality for a 40 ppb increase in 1-h max O3 concentrations for
10 lag 0-1, lag 1, and the distributed lag model (lag 0-2) was 2.70% (95% Cl: 1.02, 4.40%),
11 1.42% (95% Cl: 0.08, 2.78%), and 3.02% (95% Cl: 1.10, 4.89%), respectively. Thus, the
12 observed associations appear to occur over a short time period, (i.e., a few days).
13 Similarly, the Public Health and Air Pollution in Asia (PAPA) study (Wong etal.. 2010)
14 also examined multiple lag days (i.e., lag 0, lag 0-1, and lag 0-4), and although it did not
15 specifically examine mortality displacement it does provide additional evidence
16 regarding the timing of mortality effects proceeding O3 exposure. In a combined analysis
17 using data from all four cities examined (Bangkok, Hong Kong, Shanghai, and Wuhan),
18 excess risk estimates at lag 0-4 were larger than those at lag 0 or lag 0-1 in both fixed and
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1 random effect models (results not presented quantitatively). The larger risk estimates at
2 lag 0-4 can primarily be attributed to the strong associations observed in Bangkok and
3 Shanghai. However, it is worth noting that Bangkok differs from the three Chinese cities
4 included in this analysis in that it has a tropical climate and does not exhibit seasonal
5 patterns of mortality. As a result, Wong et al. (2010) examined the O3-mortality
6 associations at lag 0-1 in only the three Chinese cities and found that risk estimates were
7 slightly reduced from 2.26% (95% CI: 1.36, 3.16) in the 4 city analysis to 1.84% (0.77,
8 2.86) in the 3 city analysis for a 30 ppb increase in 8-h max O3 concentrations. Overall,
9 the PAPA study further supports the observation of the APHENA study that associations
10 between O3 and mortality occur over a relatively short-time period, but also indicates that
11 it may be difficult to interpret O3-mortality associations across cities with different
12 climates and mortality patterns.
13 When comparing the studies that explicitly examined the potential for mortality
14 displacement in the O3-mortality relationship, the results from Samoli et al. (2009), which
15 provide evidence that suggests mortality displacement, are not consistent with those
16 reported by Zanobetti and Schwartz (2008b). However, the shapes of the estimated
17 smooth distributed lag associations are similar (Figure 6-34 and Figure 6-35). A closer
18 examination of these figures shows that in the European data beyond a lag of 5 days the
19 estimates remain negative whereas in the U.S. data the results remain near zero for the
20 corresponding lags. These observed difference could be due the differences in the model
21 specification between the 2 studies, specifically the use of: an unconstrained distributed
22 lag model for apparent temperature up to 7 previous days (Zanobetti and Schwartz.
23 2008b) versus a linear term for humidity and an unconstrained distributed lag model of
24 temperature up to 3 previous days (Samoli et al.. 2009): and natural cubic splines with
25 2 df per season (Zanobetti and Schwartz. 2008b) versus dummy variables per month per
26 year to adjust for season (Samoli et al.. 2009). It is important to note, that these
27 differences in model specification may have also influenced the city-to-city variation in
28 risk estimates observed in these two studies (i.e., homogenous estimates across cities in
29 Zanobetti and Schwartz (2008b) and heterogeneous estimates across cities in Samoli et
30 al. (2009). Overall, the evidence of mortality displacement remains unclear, but Samoli et
31 al. (2009). Zanobetti and Schwartz (2008b). and Katsouyanni et al. (2009) all suggest that
32 the positive associations between O3 and mortality are observed mainly in the first
33 few days after exposure.
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20
Source: Reprinted with permission of BMJ Group (Samoli et al.. 2009).
The triangles represent the percent increase in all-cause mortality for a 10 |jg/m3 increase in 8-h max ozone concentrations at
each lag; the shaded area represents the 95% CIs.
Figure 6-35 Estimated combined smooth distributed lag in 21 European cities
during the summer (June-August) months.
1
2
3
4
5
6
7
8
9
10
11
12
13
Adaptation
Controlled human exposure studies have demonstrated an adaptive response to O3
exposure for respiratory effects, such as lung function decrements, but this issue has not
been examined in the epidemiologic investigation of mortality effects of O3. Zanobetti
and Schwartz (2008a) examined if there was evidence of an adaptive response in the
O3-mortality relationship in 48 U.S. cities from 1989 to 2000 (i.e., the same data analyzed
in Zanobetti and Schwartz (2008b). The authors examined all-cause mortality using a
case-crossover design to estimate the same-day (i.e., lag 0) effect of O3, matched on
referent days from every-3rd-day in the same month and year as the case. Zanobetti and
Schwartz (2008a) examined O3-mortality associations by: season, month in the summer
season (i.e., May through September), and age categories in the summer season (Table 6-
51). The estimated O3 mortality risk estimate at lag 0 was found to be highest in the
summer (1.51% [95% CI: 1.14, 1.87%]; lag 0 for a 30 ppb increase in 8-h max O3
concentrations), and, within the warm months, the association was highest in July (1.96%
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September 2011
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1
2
3
4
[95% CI: 1.42, 2.48%]; lag O).1 Upon further examination of the summer months, the
authors also observed diminished effects in August (0.84% [95% CI: 0.33, 1.39%]; lag
0). Based on these results, the authors concluded that the mortality effects of O3 appear
diminished later in the O3 season.
Table 6-51
Percent excess all-cause mortality per 10 ppb increase in daily 8-h
max ozone on the same day, by season, month, and age groups
% 95% CI
BY SEASON
Winter
Spring
Summer
Fall
-0.13
0.35
0.50
0.05
-0.56, 0.29
0.16,0.54
0.38, 0.62
-0.14,0.24
BY MONTH
May
June
July
August
September
0.48
0.46
0.65
0.28
-0.09
0.28, 0.68
0.24, 0.68
0.47, 0.82
0.11,0.46
-0.35,0.16
BY AGE GROUP
0-20
21-30
31-40
41-50
51-60
61-70
71-80
80
0.08
0.10
0.07
0.08
0.54
0.38
0.50
0.29
-0.42, 0.57
-0.67, 0.87
-0.38, 0.52
-0.27, 0.43
0.19,0.89
0.16, 0.61
0.32, 0.67
0.13,0.44
Source: Reprinted with permission from BioMed Central Ltd. (Zanobetti and Schwartz. 2008a).
5 To further evaluate the potential adaptive response observed in Zanobetti and Schwartz
6 (2008a) the distribution of the O3 concentrations across the 48 U.S. cities during July and
7 August was examined. Both July and August were found to have comparable means of
8 48.6 and 47.9 ppb with a reported maximum value of 97.9 and 96.0 ppb, respectively.
9 Thus, the observed reduction in O3-related mortality effect estimates in August (0.84%)
10 compared to July (1.96%) appears to support the existence of an adaptive response.
11 However, unlike an individual's adaptive response to decrements in lung function from
12 short-term O3 exposure, an examination of mortality prevents a direct observation of
13 adaptation. leather, for mortality the adaptation hypothesis is tested with a tacit
14 assumption that, whatever the mechanism for O3-induced mortality, the risk of death
1 These values have been standardized to the increment used throughout the ISA for max 8-h avg increase in O3 concentrations of
30 ppb. These values differ from those presented in Table 6-47 from Zanobetti and Schwartz (2008a) because the authors
presented values for a 10 ppb increase in max 8-h avg O3 concentrations.
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1 from short-term O3 exposure is reduced over the course of the summer months through
2 repeated exposures. This idea would translate to a smaller population that would die from
3 O3 exposure towards the end of summer. This may complicate the interpretation of the
4 distributed lag coefficients with long lag periods because the decreased coefficients may
5 reflect diminished effects of the late summer, rather than diminished effects that are
6 constant across the summer. These inter-twined issues need to be investigated together in
7 future research.
Ozone-Mortality Concentration-Response Relationship and Threshold
Analyses
8 Several of the recent studies evaluated have applied a variety of statistical approaches to
9 examine the shape of the O3-mortality C-R relationship and whether a threshold exists.
10 The approach used by Bell et al. (2006) consisted of applying four statistical models to
11 the NMMAPS data, which included 98 U.S. communities for the period 1987-2000.
12 These models included: a linear analysis (i.e., any change in O3 concentration can be
13 associated with mortality) (Model 1); a subset analysis (i.e., examining O3-mortality
14 relationship below a specific concentration, ranging from 5 to 60 ppb) (Model 2); a
15 threshold analysis (i.e., assuming that an association between O3 and mortality is
16 observed above a specific concentration and not below it, using the threshold values set at
17 an increment of 5 ppb between 0 to 60 ppb and evaluating a presence of a local minima in
18 AICs computed at each increment) (Model 3); and nonlinear models using natural cubic
19 splines with boundary knots placed at 0 and 80 ppb, and interior knots placed at 20 and
20 40 ppb (Model 4). A two-stage Bayesian hierarchical model was used to examine these
21 models and O3-mortality risk estimates at the city-level in the first stage analysis and
22 aggregate estimates across cities in the 2nd stage analysis using the average of 0- and
23 1-day lagged 24-h avg O3 concentrations. The results from all of these models suggest
24 that if a threshold exists it does so well below the current O3 NAAQS. When restricting
25 the analysis to all days when the current 8-h standard (i.e., 84 ppb daily 8-h max) is met
26 in each community, Bell et al. (2006) found there was still a 0.60% (95% PI: 0.30,
27 0.90%) increase in mortality per 20 ppb increase in 24-h avg O3 concentrations at lag 0-1.
28 Figure 6-36 shows the combined C-R curve obtained using the nonlinear model (Model
29 4). Although these results suggest the lack of threshold in the O3.mortality relationship, it
30 is difficult to interpret such a curve because it does not take into consideration the
31 heterogeneity in O3-mortality risk estimates across cities.
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.—
£•
o
E
o
to
re
O)
01
0-
Source: Bell et al. (2006).
Centra I estimate
95% posterior interval
0 20 40 60 80
Average of same and previous days' 0, (ppb)
Figure 6-36 Estimated combined C-R curve for ozone and nonaccidental
mortality using the nonlinear (spline) model.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
Using the same NMMAPS dataset as Bell et al. (2006). Smith et al. (20091)) further
examined the O3-mortality C-R relationship. Similar to Bell et al. (2006). Smith et al.
(2009b) conduct a subset analysis, but instead of restricting the analysis to days with O3
concentrations below a cutoff the authors only include days above a defined cutoff in the
analysis. The results of this "reversed subset" approach are in line with those reported by
Bell et al. (2006); consistent positive associations at all cutoff points up to a defined
concentration where the total number of days with O3 concentrations above a value are so
limited that the variability around the central estimate is increased. In the Smith et al.
(2009b) analysis this observation was initially observed at 45 ppb, with the largest
variability at 60 ppb; however, unlike Bell et al. (2006) where 73% of days are excluded
when subsetting the data to less than 20 ppb, the authors do not detail the number of days
of data included in the subset analyses at higher concentrations. In addition to the subset
analysis, Smith et al. (2009b) examined the shape of the C-R curve using a piecewise
linear approach with cutpoints at 40 ppb, 60 ppb, and 80 ppb. Smith et al. (2009b) found
that the shape of the C-R curve is similar to that reported by Bell et al. (2006) (Figure
6-36), but argue that slopes of the (3 for each piece of the curve are highly variable with
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1 the largest variation in the 60-80 ppb range. However, the larger variability around the (3
2 between 60-80 would be expected due to the small number of days with O3
3 concentrations within that range in an all-year analysis. This result is consistent with that
4 observed by Bell et al. (2006). which is presented in Figure 6-36.
5 The APHENA project (Katsouyanni et al.. 2009) also analyzed the Canadian and
6 European datasets (the U.S. data were analyzed for PM10 only) for evidence of a
7 threshold, using the threshold analysis method (Model 3) applied in Bell et al.'s (2006)
8 study described above. There was no evidence of a threshold in the Canadian data (i.e.,
9 the pattern of AIC values for each increment of a potential threshold value varied across
10 cities, most of which showed no local minima). Likewise, the threshold analysis
11 conducted using the European data also showed no evidence of a threshold.
12 The PAPA study, did not examine whether a threshold exists in the O3-mortality C-R
13 relationship, but instead the shape of the C-R curve individually for each city (Bangkok,
14 Hong Kong, Shanghai, and Wuhan) (Wong et al.. 2010). Using a natural spline smoother
15 with 3df for the O3 term, Wong et al. (2010) examined whether non-linearity was present
16 by testing the change in deviance between this smoothed, non-linear, model and an
17 unsmoothed, linear, model with 1 df. For each of the cities, both across the full range of
18 the O3 distribution and specifically within the range of the 25th to 75th percentile of each
19 city's O3 concentrations (i.e., a range of 9.7 ppb to 60.4 ppb across the cities) there was
20 no evidence of a non-linear relationship in the O3-mortality C-R curve. It should be noted
21 that the range of the 25th to 75th percentiles in all of the cities, except Wuhan, was lower
22 than that observed in the U.S. using all-year data where the range from the 25th to 75th
23 percentiles is 30 ppb to 50 ppb (Table 3-6).
24 Additional threshold analyses were conducted using NMMAPS data, by Xia and Tong
25 (2006) and Stylianou and Nicolich (2009). Both studies used a new statistical approach
26 developed by Xia and Tong (2006) to examine thresholds in the O3 mortality C-R
27 relationship. The approach consisted of an extended GAM model, which accounted for
28 the cumulative and nonlinear effects of air pollution using a weighted cumulative sum for
29 each pollutant, with the weights (non-increasing further into the past) derived by a
30 restricted minimization method. The authors did not use the term distributed lag model,
31 but their model has the form of distributed lag model, except that it allows for nonlinear
32 functional forms. Using NMMAPS data for 1987-1994 for 3 U.S. cities (Chicago,
33 Pittsburgh, and El Paso), Xia and Tong (2006) found that the extent of cumulative effects
34 of O3 on mortality were relatively short. While the authors also note that there was
35 evidence of a threshold effect around 24-h avg concentrations of 25 ppb, the threshold
36 values estimated in the analysis were sometimes in the range where data density was low.
37 Thus, this threshold analysis needs to be replicated in a larger number of cities to confirm
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1 this observation. It should be noted that the model used in this analysis did not include a
2 smooth function of days to adjust for unmeasured temporal confounders, and instead
3 adjusted for season using a temperature term. As a result, these results need to be viewed
4 with caution because some potential temporal confounders (e.g., influenza) do not always
5 follow seasonal patterns of temperature.
6 Stylianou and Nicolich (2009) examined the existence of thresholds following an
7 approach similar to Xia and Tong (2006) for all-cause, cardiovascular, and respiratory
8 mortality using data from NMMAPS for nine major U.S. cities (i.e., Baltimore, Chicago,
9 Dallas/Fort Worth, Los Angeles, Miami, New York, Philadelphia, Pittsburgh, and
10 Seattle) for the years 1987-2000. The authors found that PM10 and O3 were the two
11 important predictors of mortality. Stylianou and Nicolich (2009) found that the estimated
12 O3-mortality risks varied across the nine cities with the models exhibiting apparent
13 thresholds, in the 10-45 ppb range for O3. However, given the city-to-city variation in
14 risk estimates, combining the city-specific estimates into an overall estimate complicates
15 the interpretation of a threshold. Unlike the Xia and Tong (2006) analysis, Stylianou and
16 Nicolich (2009) included a smooth function of time to adjust for seasonal/temporal
17 confounding, which could explain the difference in results between the two studies.
18 In conclusion, the evaluation of the O3-mortality C-R relationship did not find any
19 evidence that supports a threshold in the relationship between short-term exposure to O3
20 and mortality within the range of O3 concentrations observed in the U.S. Additionally,
21 recent evidence suggests that the shape of the O3-mortality C-R curve remains linear
22 across the full range of O3 concentrations. However, the studies evaluated demonstrated
23 that the heterogeneity in the O3-mortality relationship across cities (or regions)
24 complicates the interpretation of a combined C-R curve and threshold analysis. Given the
25 effect modifiers identified in the mortality analyses that are also expected to vary
26 regionally (e.g., temperature, air conditioning prevalence), a national or combined
27 analysis may not be appropriate to identify whether a threshold exists in the O3-mortality
28 C-R relationship. Overall, the studies evaluated support a linear O3-mortality C-R
29 relationship and continue to support the conclusions from the 2006 O3 AQCD, which
30 stated that "if a population threshold level exists in O3 health effects, it is likely near the
31 lower limit of ambient O3 concentrations in the United States" (U.S. EPA. 2006b).
6.6.2.5 Associations of Cause-Specific Mortality and Short-term
Ozone Exposure
32 In the 2006 O3 AQCD, an evaluation of studies that examined cause-specific mortality
33 found consistent positive associations between short-term O3 exposure and
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1 cardiovascular mortality, with less consistent evidence for associations with respiratory
2 mortality. The majority of the evidence for associations between O3 exposure and cause-
3 specific mortality were from single-city studies, which had small daily mortality counts
4 and subsequently limited statistical power to detect associations.
5 New multicity studies evaluated in this review build upon and confirm the associations
6 between short-term O3 exposure and cause-specific mortality identified in the 2006 O3
7 AQCD (U.S. EPA. 2006b) (Figure 6-37; Table 6-52). In APHENA, a multicontinent
8 study that consisted of the NMMAPS, APHEA2 and Canadian multicity datasets,
9 consistent positive associations were reported for both cardiovascular and respiratory
10 mortality in all-year analyses when focusing on the natural spline model with 8 df/year
11 (Section 6.6.2.1). The associations between O3 exposure and cardiovascular and
12 respiratory mortality in all-year analyses were further supported by the multicity PAPA
13 study (Wong et al.. 2010). Cardiovascular mortality associations persisted in analyses
14 restricted to the summer season with evidence for stronger respiratory mortality
15 associations compared to those observed in all-year analyses (Figure 6-37; Table 6-52).
16 Additional multicity studies from the U.S. (Zanobetti and Schwartz. 2008b) and Europe
17 (Stafoggia et al.. 2010; Samoli et al.. 2009) that conducted summer season analyses also
18 found strong associations between O3 exposure and cardiovascular and respiratory
19 mortality.
20 Of the studies evaluated, only the APHENA study (Katsouyanni et al.. 2009) and an
21 Italian multicity study (Stafoggia et al.. 2010) conducted an analysis of the potential for
22 copollutant confounding of the O3 cause-specific mortality relationship. When focusing
23 on the natural spline model with 8 df/year and lag 1 results (as discussed in Section
24 6.6.2.1), the APHENA study found that O3 cause-specific mortality risk estimates were
25 fairly robust to the inclusion of PMi0 in copollutant models in the European dataset with
26 more variability in the U.S. and Canadian datasets (i.e., copollutant risk estimates
27 increased and decreased for respiratory and cardiovascular mortality). In summer season
28 analyses cardiovascular O3 mortality risk estimates were robust in the European dataset
29 and attenuated but remained positive in the U.S. datasets; whereas, respiratory O3
30 mortality risk estimates were attenuated in the European dataset and robust in the U.S.
31 dataset. The authors did not examine copollutant models during the summer season in the
32 Canadian dataset (Figure 6-30; Table 6-49). Interpretation of these results requires
33 caution; however, due to the different PM sampling schedules employed in each of these
34 study locations (i.e., primarily every-6th day in the U.S. and Canadian datasets and every-
35 day in the European dataset). The results of the summer season analyses from the
36 APHENA study (Katsouyanni et al.. 2009) are consistent with those from a study of 10
37 Italian cities during the summer months (Stafoggia et al.. 2010). Stafoggia et al. (2010)
38 found that cardiovascular (14.3% [95% CI: 6.7, 22.4%]) and cerebrovascular (8.5% [95%
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1
2
3
4
CI: 0.06, 16.3%]) mortality O3 effect estimates were robust to the inclusion of PM10 in
copollutant models (14.3% [95% CI: 6.7, 23.1%] and 7.3% [95% CI: -1.2, 16.3],
respectively), while respiratory mortality O3 effects estimates (17.6% [95% CI: 1.8,
35.5%]) were attenuated, but remained positive (9.2% [95% CI: -6.9, 28.8%]).
Study
Bell etal. (2005; 74345)a
Wong etal. (2010; 732535)
Katsouyanni etal. (2009; 199899)
Gryparis etal. (2004; 57276)a
Samolietal. (2009; 195855)
Zanobettiand Schwartz (2008; 101596)
Stafoggia etal. (2010; 625034)
Katsouyanni etal. (2009; 199899)
Bell etal. (2005; 74345)a
Wong etal. (2010; 732535)
Katsouyanni etal. (2009; 199899)
Gryparis etal. (2004; 57276)a
Zanobettiand Schwartz (2008; 101596)
Katsouyanni etal. (2009; 199899)
Samolietal. (2009; 195855)
Stafoggia etal. (2010; 625034)
Katsouyanni etal. (2009; 199899)
Location
U.S.andnon-U.S.
PAPA (4 cities)
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
21 European cities
21 European cities
48U.S. cities
10 Italian cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
U.S.andnon-U.S.
PAPA (4 cities)
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
21 European cities
48U.S. cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
21 European cities
10 Italian cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
Ages
All
275
<75
All
235
275
<75
All
275
All
235
275
Lag i
NR Cardiovascular ] -•- All-Year
DL(0-2)'b ]-•— ~
DL(0-2)b <-•—
DL(0-2) — | •
0-1 i — 0 — Summer
o-i i— O—
o-3 ! -O-
DL(0-2) -j O
DL(0-2)b O
DL(0-2) | O
DL(0-2)b O
NR Respiratory — •- All-Year
0-1
DL(0-2)b —
DL(0-2)b
0-1 | O Summer
0-3 — r —
DL(0-2)b ] — O —
o-i ! — c>—
DL(0-2)b ! O
-10 -5
0 5 10 15
% Increase
Effect estimates are for a 20 ppb increase in 24-h avg; 30 in 8-h max; and 40ppb increase in 1-h max ozone concentrations. Red
= cardiovascular; blue = respiratory; closed circles = all-year analysis; and open circles = summer-only analysis. An "a" represents
studies from the 2006 ozone AQCD. A "b" represents risk estimates from APHENA-Canada standardized to an approximate IQR of
5.1 ppb for a 1-h max increase in ozone concentrations (Section 6.2.7.2).
Figure 6-37 Percent increase in cause-specific mortality.
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September 2011
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Table 6-52 Corresponding effect estimates for Figure 6-37
Study
Location
Ages
Lag
Avg Time
"/.Increase (95% Cl)
Cardiovascular
All-year
Bell et al. (2005)a
Wong etal. (2010)
Katsouyanni et al. (2009)
U.S.andnon-U.S.
PAPA (4 cities)
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
All
>75
<75
NR
0-1
DL(0-2)
DL(0-2)
DL(0-2)D
DL(0-2)
DL(0-2)
DL(0-2)
DL(0-2)b
DL(0-2)
24-h avg
8-h max
1-h max
2.23(1.36,3.08)
2.20 (0.06, 4.37)
2.30 (-1.33, 6.04)
8.96(0.75,18.6)
1.1 (0.10,2.20)
2.06 (-0.24, 4.31)
3.83 (-0.1 6, 7.95)
7.03 (-2.71, 17.7)
0.87 (-0.35, 2. 10)
1.98 (-1.09, 5.1 3)
Summer
Gryparis et al. (2004)a
Samoli etal. (2009)
Zanobetti and Schwartz (2008b)
Stafoggiaetal. (2010)
Katsouyanni et al. (2009)
21 European cities
21 European cities
48 U.S. cities
10 Italian cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
All
>35
>75
<75
0-1
0-1
0-3
DL(0-5)
DL(0-2)
DL(0-2)
DL(0-2)b
DL(0-2)
DL(0-2)
DL(0-2)
DL(0-2)D
DL(0-2)
8-h max
8-h max
8-h max
8-h max
1-h max
2.7(1.29,4.32)
1.48(0.18,2.80)
2.42(1.45,3.43)
14.3(6.65,22.4)
3.1 8 (-0.47, 6.95)
1 .50 (-2.79, 5.95)
0.1 9 (-0.36, 0.74)
3.67 (0.95, 6.53)
6.78(2.70,11.0)
-1.02 (-4.23, 2.30)
-0.1 3 (-0.55, 0.29)
2.22 (-1.48, 6.04)
Respiratory
All-years
Bell et al. (2005)a
Wong etal. (2010)
Katsouyanni et al. (2009)
U.S.andnon-U.S.
PAPA (4 cities)
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
All
>75
NR
0-1
DL(0-2)
DL(0-2)
DL(0-2)b
DL(0-2)
DL(0-2)
DL(0-2)
DL(0-2)D
DL(0-2)
24-h avg
8-h max
1-h max
0.94 (-1.02, 2.96)
2.02 (-0.41, 4.49)
2.54 (-3.32, 8.79)
1.02 (-11. 9, 15.9)
0.1 3 (-1.60, 1.90)
1.82 (-2.1 8, 6.04)
1.10 (-6.48, 9.21)
-4.61 (-19.3,13.3)
-0.60 (-2.70, 1 .60)
1.10 (-3.48, 5.95)
Summer
Gryparis et al. (2004)a
Zanobetti and Schwartz (2008b)
Katsouyanni et al. (2009)
Samoli etal. (2009)
Stafoggiaetal. (2010)
Katsouyanni et al. (2009)
21 European cities
48 U.S. cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
21 European cities
10 Italian cities
APHENA-U.S.
APHENA-Canada
APHENA-Canada
APHENA-Europe
All
>35
>75
0-1
0-3
DL(0-2)
DL(0-2)
DL(0-2)D
DL(0-2)
0-1
DL(0-5)
DL(0-2)
DL(0-2)
DL(0-2)b
DL(0-2)
8-h max
8-h max
1-h max
8-h max
8-h max
1-h max
6.75(4.38,9.10)
2.51 (1.14,3.89)
4.40 (-2. 10, 11.3)
26.1(13.3,41.2)
3.00(1.60,4.50)
3.83 (-1.33, 9.21)
2.38(0.65,4.19)
17.6(1.78,35.5)
4.07 (-4.23, 13.0)
19.5(2.22,40.2)
2.30 (0.28, 4.40)
2.46 (-3.40, 8.62)
'Studies from the 2006 03 AQCD.
bRisk estimates from APHENA-Canada standardized to an approximate IQR of 5.1 ppb for a 1 -h max increase in 03 concentrations (Section
6.2.7.2).
1 Collectively, the results from the new multicity studies provide evidence of associations
2 between short-term O3 exposure and cardiovascular and respiratory mortality with
3 additional evidence indicating these associations persist, and in the case of respiratory
4 mortality are strengthened, in the summer season. Although copollutant analyses of
5 cause-specific mortality are limited, the APHENA study found that O3 cause-specific
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1 mortality risk estimates were fairly robust to the inclusion of PM10 in copollutant models
2 in the European dataset, which is supported by the results from Stafoggia et al. (2010).
3 Additionally, APHENA found that O3 cause-specific mortality risk estimates were
4 moderately to substantially sensitive (e.g., increased or attenuated) to inclusion of PMi0
5 in the U.S. and Canadian datasets. However, the mostly every-6th-day sampling schedule
6 for PMio in the U.S. and Canadian datasets greatly reduced their sample size and limits
7 the interpretation of these results.
6.6.3 Summary and Causal Determination
8 The evaluation of new multicity studies that examined the association between short-term
9 O3 exposure and mortality found evidence which supports the conclusions of the 2006 O3
10 AQCD. These new studies reported consistent positive associations between short-term
11 O3 exposure and all-cause (nonaccidental) mortality, with associations being stronger
12 during the warm season, as well as additional support for associations between O3
13 exposure and cardiovascular and respiratory mortality.
14 New studies further examined potential confounders (e.g., copollutants and seasonality)
15 of the O3-mortality relationship. Because the PM-O3 correlation varies across regions,
16 due to the difference in PM chemical constituents, interpretation of the combined effect
17 of PM on the relationship between O3 and mortality is not straightforward. Unlike
18 previous studies that were limited to primarily examining the confounding effects of
19 PM10, the new studies expanded their analyses to include multiple PM indices (e.g., PM10,
20 PM2 5, and PM components). An examination of copollutant models found evidence that
21 associations between O3 and all-cause mortality were robust to the inclusion of PM10 or
22 PM2 5 (Stafoggia etal.. 2010: Katsouvanni et al.. 2009: Bell et al.. 2007). while other
23 studies found evidence for a modest reduction (-20-30%) when examining PM10 (Smith
24 etal.. 2009b). Additional evidence suggests potential sensitivity (e.g., increases and
25 attenuation) of O3 mortality risk estimates to copollutants by age group or cause-specific
26 mortality (e.g., respiratory and cardiovascular) (Stafoggia et al.. 2010: Katsouvanni et al..
27 2009). An examination of PM components, specifically sulfate, found evidence for
28 reductions in O3-mortality risk estimates in copollutant models (Franklin and Schwartz.
29 2008). Overall, across studies, the potential impact of PM indices on O3-mortality risk
30 estimates tended to be much smaller than the variation in O3-mortality risk estimates
31 across cities suggesting that O3 effects are independent of the relationship between PM
32 and mortality. Although some studies suggest that O3-mortality risk estimates may be
33 confounded by PM or its chemical components the interpretation of these results requires
34 caution due to the limited PM datasets used as a result of the every-3rd- and 6th-day PM
35 sampling schedule. When examining the potential for seasonal confounding of the
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1 O3-mortality relationship it was observed that the extent of smoothing or the methods
2 used for adjustment can influence O3 risk estimates because of the opposing seasonal
3 trends of O3 and mortality when not instituting enough degrees of freedom to control for
4 temporal/seasonal trends (Katsouyanni et al.. 2009).
5 The multicity studies evaluated in this review also examined the regional heterogeneity
6 observed in O3-mortality risk estimates. These studies provide evidence which suggests
7 generally higher O3-mortality risk estimates in northeastern U.S. cities with some regions
8 showing no associations between O3 exposure and mortality (e.g., Southwest, Urban
9 Midwest) (Smith et al.. 2009b: Bell and Dominici. 2008). Multicity studies that examined
10 individual- and community-level characteristics identified characteristics that may
11 explain the observed regional heterogeneity in O3-mortality risk estimates as well as
12 characteristics of populations potentially susceptible to O3-related health effects. An
13 examination of community-level characteristics found an increase in the O3-mortality risk
14 estimates in cities with higher unemployment, percentage of the population
15 Black/African-American, percentage of the working population that uses public
16 transportation, lower temperatures, and lower prevalence of central air conditioning
17 (Medina-Ramon and Schwartz. 2008). Additionally, a potential interactive, or synergistic,
18 effect on the O3-mortality relationship was observed when examining differences in the
19 O3-mortality association across temperature levels (Renetal.. 2008). An examination of
20 individual-level characteristics found evidence that older age, female sex, Black race,
21 having atrial fibrillation, SES indicators (i.e., educational attainment, income level, and
22 employment status), and out-of hospital deaths, specifically in those individuals with
23 diabetes, are modify O3-mortality associations (Cakmak et al.. 2011; Stafoggia et al..
24 2010; Medina-Ramon and Schwartz. 2008). and may increase susceptibility to O3-related
25 mortality. Overall, additional research is warranted to further confirm whether these
26 characteristics, individually or in combination, can explain the observed regional
27 heterogeneity.
28 Additional studies were evaluated that examined factors, such as multi-day effects,
29 mortality displacement, adaptation, and whether a threshold exists in the O3-mortality
30 relationship, which may influence the shape of the O3-mortality C-R curve. An
31 examination of multiday effects in a U.S. and European multicity study found conflicting
32 evidence for mortality displacement, but both studies suggest that the positive
33 associations between O3 and mortality are observed mainly in the first few days after
34 exposure (Samoli et al.. 2009; Zanobetti and Schwartz. 2008b). A U.S. multicity study
35 found evidence of an adaptive response to O3 exposure, with the highest risk estimates
36 earlier in the O3 season (i.e., July) and diminished effects later (i.e., August) (Zanobetti
37 and Schwartz. 2008a). However, the evidence of adaptive effects has an implication for
38 the interpretation of multi-day effects, and requires further analysis. Analyses that
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1
2
3
4
5
6
7
9
10
11
12
13
14
15
16
specifically focused on the O3-mortality C-R relationship supported a linear O3-mortality
relationship and found no evidence of a threshold within the range of O3 concentrations
in the U.S., but did observe evidence for potential differences in the C-R relationship
across cities (Katsouyanni et al.. 2009; Stylianou and Nicolich. 2009; Bell et al.. 2006).
Collectively, these studies support the conclusions of the 2006 O3 AQCD that "if a
population threshold level exists in O3 health effects, it is likely near the lower limit of
ambient O3 concentrations in the U.S."
In conclusion, the new epidemiologic studies build upon and confirm the associations
reported in the 2006 O3 AQCD. Additionally, these new studies have provided additional
information regarding key uncertainties previously identified including the potential
confounding effects of copollutants and seasonal trend, individual- and community-level
factors that may lead to increased risk of O3-induced mortality and the heterogeneity in
O3-mortality risk estimates, and continued evidence of a linear no-threshold C-R
relationship. Although some uncertainties still remain, the collective body of evidence is
sufficient to conclude that there is likely to be a causal relationship between short-
term O3 exposure and mortality.
17
18
19
6.7 Overall Summary
The evidence reviewed in this chapter describes the recent findings regarding the health
effects of short-term exposure to ambient O3 concentrations. Table 6-53 provides an
overview of the causal determinations for each of the health categories evaluated.
Table 6-53 Summary of causal determinations for short-term exposures to
ozone
Health Category
Respiratory Effects
Cardiovascular Effects
Central Nervous System Effects
Effects on Liver and Xenobiotic Metabolism
Effects on Cutaneous and Ocular Tissues
Mortality
Causal Determination
Causal relationship
Suggestive of a causal relationship
Suggestive of a causal relationship
Inadequate to infer a causal relationship
Inadequate to infer a causal relationship
Likely to be a causal relationship
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INTEGRATED HEALTH EFFECTS OF LONG-
TERM OZONE EXPOSURE
7.1 Introduction
1 This chapter reviews, summarizes, and integrates the evidence on relationships between
2 health effects and long-term exposures to O3. Both epidemiologic and toxicological
3 studies provide a basis for examining long-term O3 exposure health effects for respiratory
4 effects, cardiovascular effects, reproductive and developmental effects, central nervous
5 system effects, cancer outcomes, and mortality. Long-term exposure has been defined as
6 a duration of approximately 30 days (1 month) or longer.
7 Conclusions from the 2006 O3 AQCD are summarized briefly at the beginning of each
8 section, and the evaluation of evidence from recent studies builds upon what was
9 available during the previous review. For each health outcome (e.g., respiratory disease,
10 lung function), results are summarized for studies from the specific scientific discipline,
11 i.e., epidemiologic and toxicological studies. The major sections (i.e. respiratory,
12 cardiovascular, mortality, reproductive/developmental, cancer) conclude with summaries
13 of the evidence for the various health outcomes within that category and integration of
14 the findings that lead to conclusions regarding causality based upon the framework
15 described in Chapter 1. Determination of causality is made for the overall health effect
16 category, such as respiratory effects, with coherence and plausibility being based on
17 evidence from across disciplines and also across the suite of related health outcomes,
18 including cause-specific mortality.
7.2 Respiratory Effects
19 Studies reviewed in the 2006 O3 AQCD examined evidence for relationships between
20 long-term O3 exposure (several months to yearly) and effects on respiratory health
21 outcomes including declines in lung function, increases in inflammation, and
22 development of asthma in children and adults. Animal toxicology data provided a clearer
23 picture indicating that long-term O3 exposure may have lasting effects. Chronic exposure
24 studies in animals have reported biochemical and morphological changes suggestive of
25 irreversible long-term O3 impacts on the lung. In contrast to supportive evidence from
26 chronic animal studies, the epidemiologic studies on longer-term (annual) lung function
27 declines, inflammation, and new asthma development remained inconclusive.
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1 Several studies (Horak et al., 2002; Frischer et al., 1999) collectively indicated that O3
2 exposure over several summer months was associated with smaller increases in lung
3 function growth in children. For longer time periods (annual), the definitive analysis in
4 the Child Health Study (CHS) reported by Gauderman et al. (2004) provided little
5 evidence that such long-term exposure to ambient O3 was associated with significant
6 deficits in the growth rate of lung function in children in contrast to the effects observed
7 with other pollutants such as acid vapor, NO2, and PM2 5. Asthmatic children with
8 GSTM1 null genotype were found to be more susceptible to the impact of O3 exposure
9 (over a 12 week study period) on small airways function in Mexico (Romieu et al..
10 2004a). Limited epidemiologic research examined the relationship between long-term O3
11 exposures and inflammation. Evidence of inflammation and allergic responses consistent
12 with known effects of O3 exposure (30 day mean) such as increased eosinophil levels
13 were observed in an Austrian study (Frischer et al., 2001). The cross-sectional surveys
14 available for the 2006 O3 AQCD detected no associations between long-term O3
15 exposures and asthma prevalence, asthma-related symptoms or allergy to common
16 aeroallergens in children after controlling for covariates.
17 New evidence presented below reports consistent associations between long-term O3
18 exposure and new-onset asthma related to genotype in U.S. cohorts in multi-community
19 studies. Related studies report coherent relationships between respiratory symptoms
20 among asthmatics and long-term O3 exposure. Short-term exposure to O3 is associated
21 with increases in respiratory symptoms and asthma medication use, as summarized in
22 Section 6.2.4.2. A new line of evidence reports a positive exposure response relationship
23 between first asthma hospitalization and long-term O3 exposure. Results from recent
24 studies examining pulmonary function, inflammation, and allergic responses are also
25 presented.
7.2.1 New Onset Asthma
26 Risk for new-onset asthma is related in part to genetic susceptibility, behavioral factors
27 and environmental exposure (Gilliland et al.. 1999). Complex chronic diseases, such as
28 asthma, are partially the result of a sequence of biochemical reactions involving
29 exposures to various environmental agents metabolized by a number of different genes
30 (Conti et al., 2003). Understanding the relation between genetic polymorphisms and
31 environmental exposure can help identify high-risk subgroups in the population and
32 provide better insight into pathway mechanisms for these complex diseases. Oxidative
33 stress likely underlies these mechanistic hypotheses (Gilliland et al.. 1999). Susceptibility
34 genes act through modification of disease risk associated with environmental factors.
35 Epidemiologic investigation of hypotheses of possible mechanisms involving the gene-
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1 environmental (GxE) interaction involves statistical analysis of these interactions for the
2 risk of new-onset asthma in children being influenced by exposure to air pollution
3 (Gauderman. 2002. 2001; Gilliland et al.. 1999).
4 Evidence for the potential importance of genetic susceptibility and behavioral factors on
5 new onset asthma are provided by several recent studies (Himes et al.. 2009; Islam et al..
6 2008; Li et al.. 2008; Hanene et al.. 2007; Ercan et al.. 2006; Li et al.. 2006a: Tamer et
7 al.. 2004; Gilliland et al.. 2002). Evidence for a gene-pollution interaction in the
8 pathogenesis of asthma are supported by recent study findings (Islam et al.. 2009; Islam
9 et al.. 2008: Orvszczvn et al.. 2007: Lee et al.. 2004b: Gilliland et al.. 2002).
10 Evidence for associations between long-term exposure to O3 and new-onset asthma is
11 provided by new studies from the CHS. Initiated in the early 1990's, the CHS was
12 originally designed to examine whether long-term exposure to ambient pollutants was
13 related to chronic respiratory outcomes in children in 12 communities in southern
14 California (Peters et al.. 1999a: Peters et al.. 1999b). About 10 years later, the CHS
15 inaugurated a series of genetic studies (Gilliland et al.. 1999) nested within the CHS
16 cohort by obtaining biological samples from the study subjects (buccal cells). These new
17 studies examined the relationship between health outcomes, genetic susceptibility,
18 behavioral factors and environmental exposure.
19 First, the hypothesis that the functional polymorphisms of HMOX-1 [(GT)n repeat], CAT
20 (-262C > T -844C > TO, and MNSOD (Ala-9Val) are associated with new-onset asthma
21 was evaluated, and then whether the effects of these variants varied by exposure to O3
22 (Islam et al.. 2008). HMOX1 [heme oxygenase (decycling) 1] is a human gene that
23 encodes for the enzyme heme oxygenase. Heme oxygenase 1 (HO-1) is an enzyme that
24 catalyzes the metabolism of heme. The heme iron serves as a source or sink of electrons
25 during electron transfer or redox chemistry, so the presence of the HMOX1 gene, and
26 therefore the generation of heme oxygenase, protects against oxidative stress in the body.
27 The authors observed that functional promoter variants in CAT and HMOX-1 showed
28 ethnicity-specific associations with new-onset asthma and that oxidant gene protection
29 was restricted to children living in low-O3 communities.
30 The subjects were drawn from the CHS cohort. Children with a history of asthma or
31 wheeze were excluded from this analysis. Analyses were restricted to children of
32 Hispanic (n = 576) or non-Hispanic white ethnicity (n = 1,125). New-onset asthma was
33 classified as having no prior history of asthma at study entry with subsequent report of
34 physician-diagnosed asthma at follow-up with the date of onset assigned to be the
35 midpoint of the interval between the interview date when asthma diagnosis was first
36 reported and the previous interview date. As a sensitivity analysis, the asthma definition
37 was restricted to those new-onset asthma cases who also used an inhaler (n = 121). They
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1 calculated long-term mean pollutant levels (1994 - 2003) to assign exposure to children
2 in each community for use in the statistical analysis. The effect of ambient air pollution
3 on the relationship between genetic polymorphism and new-onset asthma was assessed
4 using models where the community specific average air pollution levels were fitted as a
5 continuous variable together with the appropriate interaction terms for genes and air
6 pollutants (Berhane et al.. 2004). Cox proportional hazard regression models were fitted
7 to the data. A stratified analysis for the two independent fourth-grade cohorts of the study
8 population recruited in 1993 and 1996 was conducted to assess whether the results could
9 be replicated in independent groups of children.
10 Over the follow-up period, 160 new cases of asthma were diagnosed (Islam et al., 2008).
11 The evidence indicated that the effect of variation in the HMOX-1 gene on risk of new-
12 onset asthma differed by ambient O3 level. An interaction P value was reported of 0.003
13 from the hierarchical two stage Cox proportional hazard model fitting the community-
14 specific O3 and PM10 levels (continuous) and controlling for random effect of the
15 communities. Average O3 levels showed low correlation with the other monitored
16 pollutants. The interaction indicated a greater effect (association) of community O3 level
17 among children with the gene than with children without the gene. Alleles with 23 or
18 fewer (GT)n repeats are categorized as short (S). The S-allele variant of this protective
19 enzyme is more readily induced than those with more numerous repeats. The largest
20 protective effect of the (GT)n repeat polymorphism of HMOX-1 was observed for
21 children who were S-allele carriers and resided in low-O3 communities with Hazard
22 Ratio (HR) of 0.44 (95% CI: 0.23, 0.83). The ratio of HR of S-allele carriers who resided
23 in high O3 communities (HR=0.88; [95% CI: 0.33, 2.34]) was twofold greater than in
24 those who resided in the low-O3 communities (HR=0.44). The non-parallelism of the two
25 lines in An interaction p-value of 0.003 was obtained from the hierarchical two stage Cox
26 proportional hazard model fitting the community specific O3 and controlling for random
27 effect of the communities. The interaction indicates there is a greater effect (association)
28 of community O3 level on children with the gene than with children without the gene.
29 The HRs are off-set as opposed to overlapping in the figure to allow clearer presentation
30 of the results.
31 Figure 7-1 illustrates the interaction: Children with the S-allele have protection against
32 the onset of asthma; however, in high- O3 communities, this protection is attenuated. The
33 results from sensitivity analyses on the two fourth-grade cohorts, and the inhaler
34 definition for asthma were both consistent with the main results. An analysis related to
35 children's participation in sports or time spent outdoors produced the same outcome. No
36 significant interactions were observed between PM10 or other pollutants and the HMOX -
37 1 gene; quantitative results were not presented. A potential concern for not adjusting for
3 8 multiple testing was considered by the authors as not a factor in this analysis because the
Draft - Do Not Cite or Quote 7-4 September 2011
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1
2
3
selection of the genes was based on a priori hypotheses defined by a well-studied
biological pathway. Thus in this cohort in southern California, Islam et al. (2008) related
new-onset asthma to O3 exposure in genetically susceptible children.
Interaction of Gene presence and Ozone Level on the
Hazard Ratio of New Onset Asthma (P-value of 0.003)
2.5 -
2 -
1.5 -
1 -
0.5 -
0 J
(2.43)
Children with no S-Allele
0.94
•rfO.83) „ — -• **
.-. Ill ^^ r^UIMr^n ,.,^U O All^l^
Children with S-Allele
(2.34)
0.88
(0.28)
Low
(38.4 ppb)
(0.36) (0.33)
Community Mean Ozone Level
High
(55.2 ppb)
(Confidence limits based on comparison with reference group)
Source: Developed by EPA with data from Islam et al. (2008) (used by permission of American Thoracic Society).
An interaction p-value of 0.003 was obtained from the hierarchical two stage Cox proportional hazard model fitting the community
specific O3 and controlling for random effect of the communities. The interaction indicates there is a greater effect (association) of
community O3 level on children with the gene than with children without the gene. The HRs are off-set as opposed to overlapping in
the figure to allow clearer presentation of the results.
Figure 7-1 Interaction of gene presence and Os level on the Hazard Ratio (HR)
of new-onset asthma in the 12 Children's Health Study
communities.
4
5
6
7
8
9
10
11
12
13
Related to the findings in Islam et al. (2008) discussed above, Islam et al. (2009)
examined whether GSTP1, GSTM1, exercise and O3 exposure have interrelated effects
on the pathogenesis of asthma. A modifying role of air pollution on the association
between IlelOSVal and asthma in a cohort of children had been observed (Lee et al..
2004b), but the study did not examine O3 specifically or consider exercise. A primary
conclusion that the authors (Islam et al.. 2009) reported was that the common functional
variants of GSTP1 and GSTM1 null genotypes modulate the risk of new onset asthma
during adolescence. Children who had the GSTM1 null genotype were at 1.6-fold (95%
CI: 1.2, 2.2) increased risk of developing new onset asthma compared with those without
the null genotype. Further, the CHS investigators examined the complex interrelationship
Draft - Do Not Cite or Quote
7-5
September 2011
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1 of antioxidant defenses with asthma risk with increasing doses of O3, resulting from
2 increasing ventilation associated with vigorous exercise characterized by the number of
3 team sports played. In an earlier analysis, McConnell et al. (2002) had reported that the
4 risk of new onset asthma was associated with outdoor exercise, especially in high O3
5 communities but did not consider genetic variants. In this new study, Islam et al. (2009)
6 find a six fold increased risk of asthma (HR=6.15, [95% CI: 2.2, 7.4]) for children who
7 were homozygous for He 105, participated in three or more team sports and lived in
8 high-O3 communities, demonstrating the potential importance of a combination of
9 genetic variability, O3 exposure and behavior on asthma risk.
10 Epidemiologic evidence of associations of arginase variants with asthma are limited (Li
11 etal.. 2006a). Asthmatic subjects have higher arginase activity than non-asthmatic
12 subjects (Morris et al.. 2004). NO is a mediator of nitrosative stress synthesized from L-
13 arginine by nitric oxide synthases. In the CHS, Salam et al. (2009) examined whether
14 arginase variants (ARG1 and ARG2 genes) were associated with asthma and whether
15 atopy and exposures to smoking and air pollution influence the associations. The
16 modifying effect of O3 and atopy on the association between haplotypes and asthma were
17 evaluated using likelihood ratio tests with appropriate interaction terms. They found that
18 both ARG1 and ARG2 genetic loci were associated with childhood-onset asthma. The
19 effect of the ARG1 haplotype varied by the child's history of atopy and ambient O3.
20 Among atopic children living in high O3 communities, those carrying the ARG1
21 haplotype had reduced asthma risk (Odds Ratio [OR] per ARGlh4 haplotype copy =
22 0.12; [95% CI: 0.04, 0.43]; P heterogeneity across atopy/O3 categories = 0.008).
23 Further, the CHS presents results examining the relationship of new onset asthma with
24 traffic-related pollution near homes and schools (McConnell et al.. 2010). Asthma risk
25 increased with modeled traffic-related pollution exposure from roadways near homes and
26 near schools. The HR was 0.76 (95% CI: 0.38, 1.54) across the range of ambient O3
27 exposure in the communities. With adjustment for school and residential non-freeway
28 traffic-related exposure, the estimated HR for O3 was 1.01 (95% CI: 0.49, 2.11). Gene
29 variants were not evaluated in this study.
30 Some cross-sectional studies reviewed in the 2006 O3 AQCD observed positive
31 relationships between chronic exposure to O3 and prevalence of asthma and asthmatic
32 symptoms in school children (Ramadour et al.. 2000; Wang et al.. 1999) while others
33 (Kuo et al.. 2002; Charpin et al.. 1999) did not. Recent studies provide additional
34 evidence.
35 In a cross-sectional nationwide study of 32,672 Taiwanese school children, Hwang et al.
36 (2005) assessed the effects of air pollutants on the risk of asthma. The study population
37 was recruited from elementary and middle schools within 1 km of air monitoring stations.
Draft - Do Not Cite or Quote 7-6 September 2011
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1 The risk of asthma was related to O3 in the one-pollutant model. The addition of other
2 pollutants (NOX, CO2, SO2, and PMi0), in two-pollutant and three-pollutant models,
3 increased the O3 risk estimates. The prevalence of childhood asthma was assessed in
4 Portugal by contrasting the risk of asthma between a high O3 rural area and an area with
5 low O3 levels (Sousaet al., 2011; Sousa et al., 2009; Sousaet al., 2008). The locations
6 were selected to provide a difference in O3 levels without the confounding effects of
7 other pollutants. Both evaluation for asthma symptoms and FEVi suggested that O3
8 increased asthma prevalence. Clark et al. (2010) investigated the effect of exposure to
9 ambient air pollution in utero and during the first year of life on risk of subsequent
10 incidence asthma diagnosis up to 3-4 years of age in a population-based nested case-
11 control study for all children born in southwestern British Columbia in 1999 and 2000
12 (n=37,401; including 3,482 [9.3%] with asthma). Air pollution exposure for each subject
13 was estimated based on their residential address history using regulatory monitoring data,
14 land use regression modeling, and proximity to stationary pollutant sources. Daily values
15 from the three closest monitors within 50 km were used to calculate exposures. Traffic-
16 related pollutants were associated with the highest risk. Ozone was inversely correlated
17 with the primary traffic-related pollutants (r = -0.7 to -0.9). The low reliability of asthma
18 diagnosis in infants makes this study difficult to interpret (Martinez et al.. 1995). In a
19 cross-sectional analysis, Akinbami et al. (2010) examined the association between
20 chronic exposure to outdoor pollutants (12-month avg levels by county) and asthma
21 outcomes in a national sample of children ages 3-17 years living in U.S. metropolitan
22 areas (National Health Interview Survey, N = 34,073). A 5-ppb increase in estimated 8-h
23 max O3 concentration (annual average) yielded a positive association for both currently
24 having asthma and for having at least 1 asthma attack in the previous year; while the
25 adjusted odds ratios for other pollutants were not statistically significant. Models in
26 which pollutant value ranges were divided into quartiles produced comparable results.
27 Multi-pollutant models (SO2 and PM) produced similar results. The median value for
28 12-month avg O3 levels was 39.5 ppb and the IQR was 35.9-43.7 ppb. The adjusted odds
29 for current asthma for the highest quartile (49.9-59.5 ppb) of estimated O3 exposure was
30 1.56 (95% CI: 1.15,2.10) with a positive dose-response relationship apparent from the
31 lowest quartile to the highest. Thus, this cross-sectional analysis and Hwang et al. (2005)
32 provides further evidence relating O3 exposure and the risk of asthma.
33 The occurrence of bronchitic symptoms among children with asthma was investigated in
34 the CHS examining the role of gene-environment interactions and long-term O3
35 exposure. Lee et al. (2009b) studied associations of TNF-308 genotype with bronchitis
36 symptoms among asthmatic children and investigated whether associations vary with
37 ambient O3 exposure since increased airway TNF may be related to inflammation.
38 Asthmatic children with the GG genotype had a lower prevalence of bronchitic symptoms
39 compared with children carrying at least one A-allele (e.g., GA or AA). Low-versus high-
Draft - Do Not Cite or Quote 7-7 September 2011
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
O3 strata were defined as less than or greater than 50- ppb O3 avg. Asthmatic children
with TNF-308 GG genotype had a significantly reduced risk of bronchitic symptoms with
low-O3 exposure (OR=0.53; [95% CI: 0.31, 0.91]). The risk was not reduced in children
living in high-O3 communities (OR=1.42; [95% CI: 0.75, 2.70]). The difference in
genotypic effects between low- and high-O3 environments was statistically significant
among asthmatics (P for interaction = 0.01), but insignificant among non-asthmatic
children. Using indicator variables for each category based on genotype and O3 exposure,
Lee et al. (2009b) calculated the effect of TNF-308 GG genotype on the occurrence of
bronchitic symptoms among children with asthma. Figure 7-2 presents adjusted O3
community-specific beta-coefficients plotted against ambient O3 concentration, using
weights proportional to the inverse variance. They further report that they found no
substantial differences in the effect of the GG genotype in asthmatic children in relation
to exposure to PM10, PM2 5, NO2, acid vapor or second-hand smoke exposure. These
results suggest a role of gene-environment interactions such as long-term O3 exposure on
the occurrence of bronchitic symptoms among children with asthma.
Q.
2?
o
O o
O S.
03 E
o >,
CO W
£
to
20 30 40 50 60 70
Average ozone from 10 a.m. to 6 p.m. in communities (ppb)
Source: Reprinted with permission of John Wiley & Sons (Lee et al.. 2009b).
Figure 7-2 Ozone modifies the effect of TNF G-308A genotype on bronchitic
symptoms among children with asthma in the CHS. Using
indicator variables for each category based on genotype and O3
exposure, betas were calculated of TNF-308 GG genotype on the
occurrence of bronchitic symptoms among children with asthma.
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September 2011
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1 The French Epidemiology study on Genetics and Environment of Asthma (EGEA)
2 investigated the relationship between ambient air pollution and asthma severity in a
3 cohort in five French cities (Paris, Lyon, Marseille, Montpellier, and Grenoble) (Rage et
4 al.. 2009a). In this cross-sectional study, asthma severity over the past 12 months was
5 assessed among 328 adult asthmatics using two methods: (1) a four-class severity score
6 that integrated clinical events and type of treatment; and (2) a five-level asthma score
7 based only on symptoms. Two measures of exposure were also assessed: (1 [first
8 method]) closest monitor data from 1991 to 1995 where atotal of 93%ofthe subjects
9 lived within 10 km of a monitor, but where 70% of the O3 concentrations were back-
10 extrapolated values; and (2 [second method]) a validated spatial model that used
11 geostatistical interpolations and then assigned air pollutants to the geocoded residential
12 addresses of all participants and individually assigned exposure to ambient air pollution
13 estimates. Higher asthma severity scores were significantly related to both the 8-h avg O3
14 during April-September and the number of days with 8-h O3 avgs above 55 ppb. Both
15 exposure assessment methods and severity score methods resulted in very similar
16 findings. Effect estimates of O3 were similar in three-pollutant models. No PM data were
17 available. Since these estimates were not sensitive to the inclusion of ambient NO2 in the
18 three-pollutant models, the authors viewed the findings not to be explained by particles
19 which usually have substantial correlations between PM and NO2. Effect estimates for
20 O3 in three-pollutant models including O3, SO2, and NO2 yielded OR for O3-days of
21 2.74 (95% CI: 1.68, 4.48) per IQR days of 10-28 (+18) ppb. The effect estimates for SO2
22 and NO2 in the three-pollutant model were 1.33 (95% CI: 0.85, 2.11) and 0.94 (95% CI:
23 0.68, 1.29) respectively. Taking into account duration of residence did not change the
24 result. This study suggests that a higher asthma severity score is related to long-term O3
25 exposure.
26 An EGEA follow-up study (Jacquemin et al.. In Press), examines the relationship
27 between asthma and O3, NO2, and PM10. New aspects considered include: 1)
28 examination of three domains of asthma control (symptoms, exacerbations, and lung
29 function); 2) levels of asthma control (controlled, partially controlled, and uncontrolled
30 asthma); and 3) PMi0 and multi-pollutant analysis. In this cross-sectional analysis,
31 EGEA2 studied 481 adult subjects with current asthma from 2003 to 2007. The IQRs
32 were 11 (41-52) (ig/m3 for annual O3 and 13 (25-38) (ig/m3 for summer (April-
33 September) O3. The association between asthma control and air pollutants was expressed
34 by ORs (reported for one IQR of the pollutant), derived from multinomial logistic
35 regression. For each factor, the simultaneous assessment of the risk for uncontrolled
36 asthma and for partly controlled asthma was compared with controlled asthma using a
37 composite of the three domains. In crude and adjusted models, O3-sum and PM10 were
3 8 positively associated with partly controlled and uncontrolled asthma, with a clear gradient
Draft - Do Not Cite or Quote 7-9 September 2011
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1 from controlled, partly controlled (OR=1.53, 95%CI: 1.01, 2.33) and uncontrolled
2 (OR=2.14, 95% CI: 1.34, 3.43) (from the multinomial logistic regression).
3 Separately, they used a composite asthma control classification that used the ordinal
4 logistic regression for risk comparing controlled to partly controlled asthma and
5 comparing partly controlled to uncontrolled asthma. For these two pollutants, the ORs
6 assessed using the ordinal logistic regression were significant (ORs were 1.69 (95% CI:
7 1.22, 2.34) and 1.35 (95% CI: 1.13, 1.64) for O3-sum and PM10, respectively). For two
8 pollutant models using the ordinal logistic regression, the adjusted ORs for O3-sum and
9 PMio included simultaneously in a unique model were 1.50 (95% CI: 1.07, 2.11) for O3-
10 sum and 1.28 (95% CI: 1.06, 1.55) for PM10, respectively. This result suggests that the
11 effects of both pollutants are independent.
12 The analysis of the associations between air pollution for all asthma subjects and each
13 one of the three asthma control domains showed the following: 1) for lung function
14 defined dichotomously as % predicted FEVj value < or > =80 (OR=1.35, 95%CI: 0.80,
15 2.28 for adjusted O3-sum); 2) for symptoms defined as asthma attacks or dyspnoea or
16 woken by asthma attack or shortness of breath in the past three months (OR=1.59,
17 95%CI: 1.10, 2.30 for adjusted O3-sum); and for exacerbations defined at least one
18 hospitalizations or ER visits in the last year or oral corticosteroids in the past three
19 months (OR=1.58, 95%CI: 0.97, 2.59 for adjusted O3-sum). Since the estimates for both
20 pollutants were more stable and significant when using the integrated measure of asthma
21 control, this indicates that the results are not driven by one domain. These results support
22 an effect of long-term exposure to O3 on asthma control in adulthood in subjects with
23 pre-existing asthma.
24 The interrelationships between variants in catalase (CAT) and myeloperoxidase (MPO)
25 genes, ambient pollutants, and acute respiratory illness were investigated in a national
26 U.S. cohort (Wenten et al.. 2009). Health information, air pollution, and incident
27 respiratory-related school absences were ascertained in January-June 1996 for 1,136
28 Hispanic and non-Hispanic white U.S. elementary schoolchildren as part of the
29 prospective Air Pollution and Absence Study, a population based cohort study conducted
30 as part of the CHS. A related earlier study (Gilliland et al.. 2001). which was discussed in
31 the 2006 O3 AQCD, examined the effects of ambient air pollution on school absenteeism
32 due to respiratory illness without a genetic aspect to the study. In a new study Wenten et
33 al. (2009) hypothesized that variation in the level or function of these enzymes would
34 modulate respiratory illness risk, especially under high levels of oxidative stress. The
35 joint effect of these two genes on respiratory illness was examined. Risk of respiratory-
36 related school absences was elevated for children with the CAT (G/G) and MPO (G/A or
37 A/A) genes (relative risk = 1.35, [95% CI: 1.03, 1.77]; P-interaction = 0.005). To assess
Draft - Do Not Cite or Quote 7-10 September 2011
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1 effects of long-term average levels of O3 on acute effects, communities were divided into
2 high and low exposure groups by median levels (46.9 ppb O3). The epistatic effect of
3 CAT and MPO variants was evident in communities exhibiting high ambient O3 levels
4 (P-interaction = 0.03). The association of respiratory-illness absences with functional
5 variants in CAT and MPO that differ by air pollution levels illustrates the need to
6 consider genetic epistasis in assessing gene-environment interactions. In high O3
7 communities, CAT/MPO genotypes that resulted in decreased oxidative stress were
8 associated with a decreased risk of respiratory related school absences compared with the
9 CAT/MPO wild-type genotype (Relative Risk [RR] = 0.42, [95% CI: 0.20, 0.89]).
7.2.2 Asthma Hospital Admissions and ED Visits
10 The studies on O3-related hospital discharges and emergency department (ED) visits for
11 asthma and respiratory disease that were available in the 2006 O3 AQCD mainly looked
12 at the daily time metric. Collectively the short-term O3 studies presented earlier in
13 Section 6.2.7.5 indicate that there is evidence for increases in both hospital admissions
14 and ED visits related to both all respiratory outcomes and asthma with stronger
15 associations in the warm months. New studies evaluated long-term O3 exposure metrics
16 providing a new line of evidence that suggests a positive exposure-response relationship
17 between first asthma hospital admission and long-term O3 exposure.
18 An ecologic study (Moore et al.. 2008) evaluated time trends in associations between
19 declining warm-season O3 concentrations and hospitalization for asthma in children in
20 California's South Coast Air Basin who ranged in age from birth to 19 years. Quarterly
21 average concentrations from 195 spatial grids, 10* 10 km, were used. Ozone was the only
22 pollutant associated with increased hospital admissions over the study period. A linear
23 relation was observed for asthma hospital discharges (Moore et al.. 2008). A matched
24 case-control study (Karr et al.. 2007) was conducted of infant bronchiolitis (ICD 9, code
25 466.1) hospitalization and two measures of long-term pollutant exposure (the month prior
26 to hospitalization and the lifetime average) for O3 in the South Coast Air Basin of
27 southern California among 18,595 infants born between 1995 and 2000. Ozone was
28 associated with reduced risk in the single-pollutant model, but this relation did not persist
29 in multi-pollutant models (CO, NO2 and PM2 5).
30 In a cross-sectional study, Meng et al. (2010) examined associations between air
31 pollution and asthma morbidity in the San Joaquin Valley in California by using the 2001
32 California Health Interview Survey data from subjects ages 1 to 65+ who reported
33 physician-diagnosed asthma (n = 1,502). Subjects were assigned annual average
34 concentrations for O3 based on residential ZIP code and the closet air monitoring station
Draft - Do Not Cite or Quote 7-11 September 2011
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1 within 8 km but did not have data on duration of residence. Multi-pollutant models for O3
2 and PM did not differ substantially from single-pollutant estimates, indicating that
3 pollutant multi-collinearity is not a problem in these analyses. The authors reported
4 increased asthma-related ED visits or hospitalizations for O3 (OR=1.49; [95% CI: 1.05,
5 2.11] per 10 ppb) for all ages. Positive associations were obtained for symptoms but 95%
6 CIs included null values. Associations for symptoms for adults (ages 18 +) were observed
7 (OR=1.40; [95% CI: 1.02, 1.91] per 10 ppb).
8 Associations between air pollution and poorly controlled asthma among adults in
9 Los Angeles and San Diego Counties, were investigated using the California Health
10 Interview Survey data collected between November 2000 and September 2001 (Meng et
11 al.. 2007). Each respondent was assigned an annual average concentration measured at
12 the nearest station within 5 miles of the residential cross-street intersection. Poorly
13 controlled asthma was defined as having daily or weekly asthma symptoms or at least one
14 ED visit or hospitalization because of asthma during the past 12 months. This cross-
15 sectional study reports an OR of 3.34 (95% CI: 1.01, 11.09) for poorly controlled asthma
16 when comparing those 65 years of age and older above the 90th percentile (28.7 ppb)
17 level to those below that level. Multi-pollutant (PM) analysis produced similar results.
18 Evidence associating long-term O3 exposure to first asthma hospital admission in a
19 concentration-response relationship is provided in a retrospective cohort study (Lin et al..
20 2008b). This study investigated the association between chronic exposure to O3 and
21 childhood asthma admissions (defined as a principal diagnosis of ICD9, code 493) by
22 following a birth cohort of 1,204,396 eligible births born in New York State during 1995-
23 1999 to first asthma admission or until 31 December 2000. There were 10,429 (0.87%)
24 children admitted to the hospital for asthma between 1 and 6 years of age. The asthma
25 hospitalization rate in New York State in 1993 was 2.87 per 1,000 (Linetal.. 1999).
26 Three annual indicators (all 8-h max from 10:00 a.m. to 6:00 p.m.) were used to define
27 chronic O3 exposure: (1) mean concentration during the follow-up period (41.06 ppb); (2)
28 mean concentration during the O3 season (50.62 ppb); and (3) proportion of follow-up
29 days with O3 levels >70 ppb. In this study the authors aimed to predict the risk of having
30 asthma admissions in a birth cohort, but the time to the first admission in children that is
31 usually analyzed in survival models was not their primary interest. The effects of
32 co-pollutants were assessed and controlled for using the Air Quality Index (AQI).
33 Interaction terms were used to assess potential effect modifications. A positive
34 association between chronic exposure to O3 and childhood asthma hospital admissions
35 was observed indicating that children exposed to high O3 levels over time are more likely
36 to develop asthma severe enough to be admitted to the hospital. The various factors were
37 examined and differences were found for younger children (1-2 years), poor
38 neighborhoods, Medicaid/self-paid births, geographic region and others. As shown in
Draft - Do Not Cite or Quote 7-12 September 2011
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1
2
3
4
5
6
9
10
11
12
13
14
15
16
Adjusted for child's sex, age, birth weight, and gestational age; maternal race, ethnicity, age, education,
insurance, and smoking status during pregnancy; and regional poverty level and temperature. ORs by low,
medium, and high exposure are shown for New York City (NYC: low [37.3 ppb],
medium [37.3 -38.11] ppb, high [38.11 + ppb]) and other New York State regions (Other
NYS regions: low [42.58 ppb], medium [42.58-45.06 ppb], high [45.06+ ppb]) for first
asthma hospital admission.
Figure 7-3, positive concentration-response relationships were observed. Asthma
admissions were significantly associated with increased O3 levels for all chronic
exposure indicators (ORs, 1.16-1.68). When estimating the O3 effect using the
exceedance proportion, an increase was observed (OR=1.68; [95% CI: 1.64, 1.73]) in
hospital admissions with an IQR (2.51%) increase in O3. A proportional hazards model
for the New York City data was run as a sensitivity analysis and it yielded similar results
between asthma admissions and chronic exposure to O3 (Cox model: HR = 1.14, [95%
CI: 1.124, 1.155] is similar to logistic model results: OR= 1.16 (95% CI: 1.15, 1.17)
(Lin. 2010). Thus, this study provides evidence associating long-term O3 exposure to first
asthma hospital admission in a concentration-response relationship.
3.0
2.5
« 2.0
^p
o^
8 1.5
g 1.0
0.5
n
i i Low exposure 0-33%
i i Medium exposure 34-66%
•zzi High exposure ^ 67%
1
(1.29
1.00
(ref)
43
-1.58)
T
1
1.6E
1.52-1
1
.80) (1.4E
1.00
(ref)
(1
.64
-1.82)
T
1
2.06
07 9 n
7)
New York City
Other NYS regions
Regions
Adjusted for child's sex, age, birth weight, and gestational age; maternal race, ethnicity, age, education, insurance, and smoking
status during pregnancy; and regional poverty level and temperature. ORs by low, medium, and high exposure are shown for
New York City (NYC: low [37.3 ppb], medium [37.3 - 38.11] ppb, high [38.11 + ppb]) and other New York State regions (Other NYS
regions: low [42.58 ppb], medium [42.58-45.06 ppb], high [45.06+ ppb]) for first asthma hospital admission.
Figure 7-3 Ozone-asthma concentration-response relationship using the mean
concentration during the entire follow-up period.
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September 2011
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7.2.3 Pulmonary Structure and Function
1 The definitive 8-year follow-up analysis of the first cohort of the CHS, which is
2 discussed in Section 7.2 (Gauderman et al., 2004). provided little evidence that long-term
3 exposure to ambient O3 was associated with significant deficits in the growth rate of lung
4 function in children. A later CHS study (Islam et al., 2007) examined relationships
5 between air pollution, lung function, and new onset asthma and reported no substantial
6 differences in the effect of O3 on lung function. Ozone concentrations from the least to
7 most polluted communities (mean annual average of 8-h avg O3) ranged from 30 to
8 65 ppb, as compared to the ranges observed for the other pollutants, which had four- to
9 eightfold differences in concentrations. In a more recent CHS study, Breton et al. (2011)
10 hypothesized that genetic variation in genes on the glutathione metabolic pathway may
11 influence the association between ambient air pollutant exposures and lung function
12 growth in children. They investigated whether genetic variation in glutathione genes
13 GSS, GSR, GCLC, and GCLM was associated with lung function growth in healthy
14 children using data collected on 2,106 children over an 8-year time-period as part of the
15 Children's Health Study. Breton et al. (2011) found that variation in the GSS locus was
16 associated with differences in susceptibility of children for lung function growth deficits
17 associated with NO2, PMi0, PM25, elemental carbon, organic carbon, and O3. The
18 negative effects of air pollutants were largely observed within participants who had a
19 particular GSS haplotype. The effects ranged from -124.2 to -149.1 mL for FEVi, -92.9
20 to -126.7 mL for FVC and -193.9 to -277.9 mL/s for MMEF for all pollutants except O3,
21 for which some positive associations were reported: 25.9 mL for FEVi; 0.1 mL for FVC,
22 and 166.5 mL/s for MMEF. Ozone was associated with larger decreases in lung function
23 in children without this haplotype, when compared to the other pollutants with values of -
24 76.6 mL for FEVj, -17.2 mL for FVC, and -200.3 mL/s for MMEF, but only the
25 association with MMEF was statistically significant.
26 As discussed in the 2006 O3 AQCD, a study of freshman students at the University of
27 California, Berkeley reported that lifetime exposure to O3 was associated with decreased
28 measures of small airways (<2 mm) function (FEF75 and FEF25_75) (Tager et al.. 2005).
29 There was an interaction with the FEF25-75/FVC ratio, a measure of intrinsic airway size.
30 Subjects with a large ratio were less likely to have decreases in FEF75 and FEF25_75 for a
31 given estimated lifetime exposure to O3. Kinney and Lippmann (2000) examined 72
32 nonsmoking adults (mean age 20 years) from the second-year class of students at the U.S.
33 Military Academy in West Point, NY, and reported results that appear to be consistent
34 with a decline in lung function that may in part be due to O3 exposures over a period of
35 several summer months. Ilhorst et al. (2004) examined 2,153 children with a median age
36 of 7.6 years and reported pulmonary function results which indicated that significantly
37 lower FVC and FEVi increases were associated with higher O3 exposures over the
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1 medium-term of several summer months, but not over several months in the winter.
2 Semi-annual mean O3 concentrations ranged from 22 to 54 ppb during the summer
3 months and 4 to 36 ppb during the winter months. However, over the longer-term
4 3.5-year period Ilhorst et al. (2004) found no associations between increases in lung
5 function and mean summer months O3 levels for FVC and FEVi, in contrast to the
6 significant medium-term effects. Frischer et al. (1999) showed results similar to the
7 Ilhorst et al. (2004) study.
8 Mortimer et al. (2008a, b) examined the association of prenatal and lifetime exposures to
9 air pollutants with pulmonary function and allergen sensitization in a subset of asthmatic
10 children (ages 6-11) included in the Fresno Asthmatic Children's Environment Study
11 (FACES). Monthly means of pollutant levels for the years 1989-2000 were created and
12 averaged separately across several important developmental time-periods, including: the
13 entire pregnancy, each trimester, the first 3 years of life, the first 6 years of life, and the
14 entire lifetime. In the first analysis (Mortimer et al.. 2008a). negative effects on
15 pulmonary function were found for exposure to PMi0, NO2, and CO during key neonatal
16 and early life developmental periods. The authors did not find a negative effect of
17 exposure to O3 within this cohort. In the second analysis (Mortimer et al.. 2008b).
18 sensitization to at least one allergen was associated, in general, with higher levels of CO
19 and PM10 during the entire pregnancy and second trimester, and higher PM10 during the
20 first 2 years of life. Lower exposure to O3 during the entire pregnancy or second trimester
21 was associated with an increased risk of allergen sensitization. Although the pollutant
22 metrics across time periods were correlated, the strongest associations with the outcomes
23 were observed for prenatal exposures. Though it may be difficult to disentangle the effect
24 of prenatal and postnatal exposures, the models from this group of studies suggest that
25 each time period of exposure may contribute independently to different dimensions of
26 school-aged children's pulmonary function. For 4 of the 8 pulmonary-function measures
27 (FVC, FEVi, PEF, FEF25.75), prenatal exposures were more influential on pulmonary
28 function than early-lifetime metrics, while, in contrast, the ratio of measures (FEVi/FVC
29 and FEF25_75/FVC) were most influenced by postnatal exposures. When lifetime metrics
30 were considered alone, or in combination with the prenatal metrics, the lifetime measures
31 were not associated with any of the outcomes. This suggests that the timing of the O3
32 exposure may be more important than the overall dose, and prenatal exposures are not
33 just markers for lifetime or current exposures.
34 Latzin et al. (2009) examined whether prenatal exposure to air pollution was associated
35 with lung function changes in the newborn. Tidal breathing, lung volume, ventilation
36 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age =
37 5 weeks). Consistent with the previous studies, no association was found for prenatal
38 exposure to O3 and lung function.
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1 In a cross-sectional study of adults, Qian et al. (2005) examined the association oflong-
2 term exposure to O3 and PMi0 with pulmonary function from data of 10,240 middle-aged
3 subjects who participated in the Atherosclerosis Risk in Communities (ARIC) study in
4 four U.S. communities. A surrogate for long-term O3 exposure from daily data was
5 determined at the individual level. Ozone was significantly and negatively associated
6 with measures of pulmonary function.
7 To determine the extent to which long-term exposure to outdoor air pollution accelerates
8 adult decline in lung function, Forbes et al. (2009b) studied the association between
9 chronic exposure to outdoor air pollution and lung function in approximately 42,000
10 adults aged 16 and older who were representatively sampled cross-sectionally from
11 participants in the Health Survey for England (1995, 1996, 1997, and 2001). FEVj was
12 not associated with O3 concentrations. In contrast to the results for PM10, NO2, and SO2
13 combining the results of all the survey years showed that a 5-ppb difference in O3 was
14 counter-intuitively associated with a higher FEVi by 22 mL.
15 In a prospective cohort study consisting of school-age, non-asthmatic children in
16 Mexico City (n = 3,170) who were 8 years of age at the beginning of the study, Rojas-
17 Martinez et al. (2007) evaluated the association between long-term exposure to O3, PM10
18 and NO2 and lung function growth every 6 months from April 1996 through May 1999.
19 Exposure data were provided by 10 air quality monitor stations located within 2 km of
20 each child's school. Over the study period, 8-h O3 concentrations ranged from 60 ppb
21 (SD, ±25) in the northeast area of Mexico City to 90 ppb (SD, ±34) in the southwest, with
22 an overall mean of 69.8 ppb. In multi-pollutant models, an IQR increase in mean O3
23 concentration of 11.3 ppb was associated with an annual deficit in FEVi of 12 mL in
24 girls and 4 mL in boys. Single-pollutant models showed an association between ambient
25 pollutants (O3, PMi0 and NO2) and deficits in lung function growth. While the estimates
26 from copollutant models were not substantially different than single pollutant models,
27 independent effects for pollutants could not be estimated accurately because the traffic -
28 related pollutants were correlated. To reduce exposure misclassification,
29 microenvironmental and personal exposure assessments were conducted in a randomly
30 selected subsample of 60 children using passive O3 samplers. Personal O3 concentrations
31 were correlated (p < 0.05) with the measurements obtained from the fixed-site air
32 monitoring stations.
33 In the 2006 O3 AQCD, few studies had investigated the effect of chronic O3 exposure on
34 pulmonary function. The strongest evidence was for medium-term effects of extended O3
35 exposures over several summer months on lung function in children, i.e., reduced lung
36 function growth being associated with higher ambient O3 levels. Longer-term studies,
37 investigating the association of chronic O3 exposure on lung function such as the CHS,
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1 were inconclusive. Short-term O3 exposure studies presented in Section 6.2.1.2 provide a
2 cumulative body of epidemiologic evidence that strongly supports associations between
3 ambient O3 exposure and decrements in lung function among children. For new studies
4 of long-term O3 exposure relationship to pulmonary function, one study, where O3 and
5 other pollutant levels were higher (90 ppb at high end of the range) than those in the
6 CHS, observes a relationship between O3 concentration and pulmonary function declines
7 in school-aged children. Two studies of adult cohorts provide mixed results where long-
8 term exposures were at the high end of the range with levels of 49.5 ppb in one study and
9 27 ppb IQR in the other. Thus there is little new evidence to build upon the very limited
10 studies from the 2006 O3 AQCD.
7.2.3.1 Pulmonary Structure and Function: Evidence from
lexicological Studies
11 As reviewed in the 1996 and 2006 O3 AQCDs, there are both qualitative and quantitative
12 uncertainties in the extrapolation of data generated by rodent toxicology studies to the
13 understanding of health effects observed in humans, as documented by epidemiologic and
14 controlled exposure studies. Chief among these data extrapolation issues are the
15 differences between rodent and human respiratory physiology, cellular makeup,
16 dosimetry, and morphometry (see Chapter 5). However, important evidence is available
17 from O3-inhalation studies performed in nonhuman primates whose respiratory system
18 most closely resembles that of the human. A long series of studies have used nonhuman
19 primates to examine the effect of O3 alone or in combination with an inhaled allergen,
20 house dust mite antigen, on morphology and lung function. These studies, by Plopper and
21 colleagues, have demonstrated changes in pulmonary function and airway morphology in
22 adult and infant nonhuman primates repeatedly exposed to environmentally relevant
23 concentrations of O3 (Joad et al.. 2008; Carey et al.. 2007; Plopper et al.. 2007; Fanucchi
24 et al.. 2006; Joad et al.. 2006; Evans et al.. 2004; Larson et al.. 2004; Tran et al.. 2004;
25 Evans et al.. 2003b; Schelegle etal.. 2003: Fanucchi et al.. 2000; Hydeetal.. 1989;
26 Harkema et al.. 1987a; Harkema et al.. 1987b; Fuiinaka et al.. 1985). The findings of
27 these nonhuman primate studies have also been observed in rodent studies discussed at
28 the end of this section and included in Table 7-1.
29 Since publication of the 1996 and 2006 O3 AQCDs, the initial observations in adult
30 nonhuman primates have been expanded in a series of experiments using infant rhesus
31 monkeys repeatedly exposed to 0.5 ppm O3 starting at 1 month of age (Plopper et al.,
32 2007). Many of the observations found in adult monkeys have also been noted in infant
33 rhesus monkeys, although a direct comparison of the degree of effects between adult and
34 infant monkeys has not been reported. In terms of pulmonary function changes, after
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1 several episodic exposures of infant monkeys to O3 (each cycle: 5 days of 0.5 ppm O3 at
2 8 h/day, followed by 9 days of filtered air exposure), they observed more than a doubling
3 in the baseline airway resistance, which was accompanied by a small increase in airway
4 responsiveness to inhaled histamine (Schelegle et al.. 2003). although neither
5 measurement was statistically different from filtered air control values. Exposure of
6 animals to inhaled house dust mite antigen alone also produced small but not statistically
7 significant changes in baseline airway resistance and airway responsiveness, whereas the
8 combined exposure to both (O3 + antigen) produced statistically significant and greater
9 than additive changes in both functional measurements. This nonhuman primate evidence
10 of an O 3-induced change in airway responsiveness supports the biologic plausibility of
11 long-term exposure to O3 contributing to the effects of asthma in children. To understand
12 which conducting airways and inflammatory mechanisms are involved in O3-induced
13 airway hyperresponsiveness in the infant rhesus monkey, a follow-up study examined
14 airway responsiveness ex vivo in lung slices (Joad et al.. 2006). Using video microscopy
15 to morphometrically evaluate the response of bronchi and respiratory bronchioles to
16 methacholine, (a bronchoconstricting agent commonly used to evaluate airway
17 responsiveness in asthmatics), the investigators observed differential effects for the two
18 airway sizes. While episodic exposure to O3 alone (0.5 ppm) had little effect on ex vivo
19 airway responsiveness in bronchi and respiratory bronchioles, exposure to dust mite
20 antigen alone produced airway hyperresponsiveness in the large bronchi, whereas O3 +
21 antigen produced significant increases in airway hyperresponsiveness only in the
22 respiratory bronchioles. These results suggest that ozone's effect on airway
23 responsiveness occurs predominantly in the smaller bronchioles, where dosimetric
24 models indicate the dose would be higher.
25 The functional changes in the conducting airways of infant rhesus monkeys exposed to
26 either O3 alone or O3 + antigen were accompanied by a number of cellular and
27 morphological changes, including a significant fourfold increase in eosinophils, (a cell
28 type important in allergic asthma), in the bronchoalveolar lavage of infant monkeys
29 exposed to O3 alone. Thus, these studies demonstrate both functional and cellular
30 changes in the lung of infant monkeys after cyclic exposure to 0.5 ppm O3. This
31 concentration, while higher than those used in controlled human exposure studies,
32 provides relevant information to understanding the adverse effects of ambient O3
33 exposure on the respiratory tract of humans. No concentration-response data, however,
34 are available from these nonhuman primate studies.
35 In addition to these functional and cellular changes, significant structural changes in the
36 respiratory tract have been observed in infant rhesus monkeys exposed to O3. During
37 normal respiratory tract development, conducting airways increase in diameter and length
38 in the infant rhesus monkey. Exposure to O3 alone (5 days of 0.5 ppm O3 at 8 h/day,
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1 followed by 9 days of filtered air exposures for 11 cycles), however, markedly affected
2 the growth pattern of distal conducting airways (Fanucchi et al.. 2006). Whereas the first
3 alveolar outpocketing occurred at airway generation 13 or 14 in filtered air-control infant
4 monkeys, the most proximal alveolarized airways occurred at an average of 10 airway
5 generations in O3-exposed monkeys. Similarly, the diameter and length of the terminal
6 and respiratory bronchioles were significantly decreased in O3-exposed monkeys.
7 Importantly, the O3-induced structural pathway changes persisted after recovery in
8 filtered air for 6 months after cessation of the O3 exposures. These structural effects were
9 accompanied by significant increases in mucus goblet cell mass, alterations in smooth
10 muscle orientation in the respiratory bronchioles, epithelial nerve fiber distribution, and
11 basement membrane zone morphometry. These latter effects are significant because of
12 their potential contribution to airway obstruction and airway hyperresponsiveness which
13 are central features of asthma.
14 As noted above, a significant increase in airway responsiveness to inhaled histamine
15 occurred in infant rhesus monkeys exposed to O3 + house dust mite antigen, but not to O3
16 alone (Schelegle et al.. 2003). To study the underlying mechanisms of this airway
17 hyperresponsiveness, these investigators evaluated the effect of exposure to O3 alone and
18 in combination with (+) antigen on non-specific airway responsiveness to methacholine
19 at different airway generations. After exposure to filtered air, O3, antigen, or O3 +
20 antigen, the bronchi and respiratory bronchioles of 6-month-old monkeys were
21 challenged ex vivo with methacholine. Exposure to O3 alone had no significant effect on
22 airway responsiveness to methacholine in either airway, whereas O3 + antigen produced
23 a significant increase in airway responsiveness in the respiratory bronchioles but not the
24 larger bronchi.
25 Because many cellular and biochemical factors are known to contribute to allergic
26 asthma, the effect of exposure to O3 alone or O3 + antigen on immune system parameters
27 was also examined in infant rhesus monkeys. Mast cells, which contribute to asthma via
28 the release of potent proteases, were elevated in animals exposed to antigen alone but O3
29 alone had little effect on mast cell numbers and the response of animals exposed to O3 +
30 antigen was not different from that of animals exposed to antigen alone; thus suggesting
31 that mast cells played little role in the interaction between O3 and antigen in this model of
32 allergic asthma (Van Winkle et al.. 2010). Increases in CD4+ and CD8+ lymphocytes
33 were observed at 6 months of age in the blood and bronchoalveolar lavage fluid of infant
34 rhesus monkeys exposed to O3 + antigen but not in monkeys exposed to either agent
35 alone (Miller et al.. 2009). Activated lymphocytes (i.e., CD25+ cells) were
36 morphometrically evaluated in the airway mucosa and significantly increased in infant
37 monkeys exposed to antigen alone or O3 + antigen. Although O3 alone had no effect on
3 8 CD25+ cells, it did alter the anatomic distribution of CD25+ cells within the airways.
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1 Ozone had only a small effect on these sets of immune cells and did not produce a strong
2 interaction with an inhaled allergen in this nonhuman primate model.
3 In addition to alterations in the immune system, nervous system interactions with
4 epithelial cells are thought to play a contributing role to airway hyperresponsiveness. As
5 noted in the 2006 O3 AQCD, exposure of infant rhesus monkeys altered the normal
6 development of neural innervation in the epithelium of the conducting airways (Larson et
7 al.. 2004). Whereas, a significant reduction in airway innervation occurred after exposure
8 to O3 alone, a significantly greater reduction occurred in monkeys exposed to O3 +
9 antigen. This reduction in overall airway innervation was accompanied, however, by an
10 increase in the abundance of protein gene product 9.5, a nonspecific neural marker.
11 Significant increases in protein gene product 9.5 were still observed in O3 alone- and O3
12 + antigen-exposed infant monkeys after a 6-month recovery protocol (Kajekar et al.,
13 2007). Thus, in addition to structural, immune, and inflammatory effects, exposure to O3
14 produces alterations in airway innervation which may contribute to O3-induced
15 exacerbation of asthma.
16 A number of studies in both nonhuman primates and rodents demonstrate that O3
17 exposure can increase collagen synthesis and deposition, inducing fibrotic-like changes in
18 the lung (Lastetal.. 1994; Chang etal.. 1992; Moffatt et al.. 1987; Reiser et al.. 1987;
19 Lastetal.. 1984). Increased collagen content is often associated with elevated abnormal
20 cross links that appear to be irreversible (Reiser et al.. 1987). Generally changes in
21 collagen content have been observed in rats exposed to 0.5 ppm O3 or higher, although
22 extracellular matrix thickening has been observed in the lungs of rats exposed to an urban
23 pattern of O3 with daily peaks of 0.25 ppm for 38 weeks (Chang etal.. 1992; Chang et
24 al., 1991). A more recent study using an urban pattern of exposure to 0.5 ppm O3
25 demonstrated that O3-induced collagen deposition in mice is dependent on the activity of
26 TGF-(3 (Katre etal.. 2011). Sex differences have been observed with respect to increased
27 centriacinar collagen deposition and crosslinking, which was observed in female but not
28 male rats exposed to 0.5 and 1.0 ppm O3 for 20 months (Lastetal.. 1994). Few other
29 long-term exposure morphological studies have presented sex differences and most only
30 evaluated males. It is unclear what the long-term effects of these structural changes may
31 be. A number of studies indicate that structural changes in the respiratory system are
32 persistent or irreversible. For example, O3-induced hyperplasia was still evident in the
33 nasal epithelia of rats 13 weeks after recovery from 0.5 ppm O3 exposure (Harkema et
34 al., 1999). In a study of episodic exposure to 0.25 ppm O3, Chang et al. (1992) observed
35 no reversal of basement membrane thickening in rat lungs up to 17 weeks post-exposure.
36 Episodic exposure (0.25 ppm O3, every other month) of monkeys induced equivalent
37 morphological changes compared to continuously exposed animals, even though they
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1 were exposed for half the time and evaluation occurred a month after exposure ceased as
2 opposed to immediately (Tyler etal.. 1988).
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Table 7-1 Respiratory effects in nonhuman primates and rodents resulting
from long-term Oa exposure
Study
Catalanoetal. (1995a;
1995b); Chang et al.
(1995): Harkema et al.
(1997a:1997b:1994):
Last etal. (1994);
Pinkertonetal. (1995);
Plopperetal.(1994);
Stockstill et al. (1995);
Pinkertonetal. (1998)
Herbert etal. (1996)
Chang et al. (1991)
Chang et al. (1992)
Barry etal. (1985. 1983)
Tyler etal. (1988)
Harkema et al. (1999)
Van Bree etal. (2002)
Katre etal. (2011)
Model
Rat, male and
female, Fischer
F344, 6-8 weeks
old
Mice, male and
female, B6C3F1,
6-7 weeks old,
Rat, male, F344, 6
weeks old
Rat, male, F344, 6
weeks old
Rat, male, 1 day
old or 6 weeks old
Monkey; male,
Macaca
fascicularis, 7 mo
old
Rat, male, Fischer
F344/N HSD, 10-
14 weeks old
Rat, male, Wistar,
7 weeks old, n =
5/group
Mice; male,
C57BL/6,
6-8 week sold
O3 (ppm)
0.12
0.5
1.0
0.12
0.50
1.0
Continuous: 0.1 2
or 0.25
Episodic/urban:
baseline 0.06;
peak 0.25
baseline 0.06;
peak 0.25
0.1 2 (adults only)
0.25
0.25
0.25
0.5
0.4
0.5
Exposure Duration
6 h/day, 5 days/week for
20 months
6 h/day, 5 days/week for
24 and 30 months
Continuous: 12 h/day for
6 weeks
Simulated urban pattern;
slow rise to peak 9 h/day,
5 days/week, 13 weeks
Slow rise to peak
9 h/day, 5 days/week,
13 and 78 weeks
Recovery in filtered
air for 6 or 17 weeks
12 h/day for 6 weeks
8 h/day, 7 days/week,
Daily for 18 mo or
episodically every other
mo for 18mos
Episodic group
evaluated 1 mo post
exposure
8 h/day, 7 days/week for
13 weeks
23.5 h/day for 1,3,7,
28,or 56 days
8 h/day, [5 days/week
03, and 2 days filtered
air] for 5 or 10 cycles
Effects
Effects similar to (or a model of) early fibrotic human
disease were greater in the periacinar region than in
terminal bronchioles. Thickened alveolar septa
observed in rats exposed to 0.12 ppm 03. Other
effects (e.g., mucous cell metaplasia in the nose and
mild fibrotic response in the parenchyma, increased
collagen in CAR of females) observed at 0.5 to 1 .0
ppm. Some morphometric changes such as epithelial
thickening and bronchiolarization occurred after 2 or 3
months of exposure to 1 .0 ppm.
Similar to the response of rats in the same study (see
rat above). Effects were seen in the nose and
centriacinar region of the lung at 0.5 and 1 .0 ppm.
Increased Type 1 and 2 epithelial volume assessed by
TEM. Linear relationship observed between increases
in Type 1 epithelial cell volume and concentration x
time product. Degree of injury not related to pattern of
exposure (continuous or episodic).
Progressive epithelial hyperplasia, fibroblast
proliferation, and interstitial matrix accumulation
observed using TEM. Interstitial matrix thickening due
to deposition of basement membrane and collagen
fibers. Partial recovery of interstitial matrix during
follow-up periods in air; but no resolution of basement
membrane thickening.
Lung and alveolar development not significantly
affected. Increased Type 1 and 2 epithelial cells and
AM in CAR alveoli, thickened Type 1 cells with smaller
volume and less surface coverage as assessed by
TEM (adults and juveniles). In adults, smaller but
statistically significant similar changes at 0.12 ppm,
suggesting linear concentration-response relationship.
No statistically significant age-related effects
observed.
Increased collagen content, chest wall compliance,
and inspiratory capacity in episodic group only.
Respiratory bronchiolitis in both groups. Episodically
exposed group incurred greater alterations in
physiology and biochemistry and equivalent changes
in morphometry even though exposed for half the time
as the daily exposure group.
Mucous cell hyperplasia in nasal epithelium after
exposure to 0.25 and 0.5 ppm 03; still evident after 13
weeks recovery from 0.5 ppm 03 exposure.
Acute inflammatory response in BALF reached a
maximum at day 1 and resolved within 6 days during
exposure. Centriacinar region inflammatory responses
throughout 0 3 exposure with increased collagen and
bronchiolization still present after a recovery period.
Sustained elevation in TGF-p and PAI-1 in lung (5 or
10 cycles); elevated a-SMA and increased collagen
deposition in airway walls (after 10 cycles). Collagen
increase shown to depend on TGF-p.
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September 2011
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Study
Model
(ppm) Exposure Duration
Effects
Schelegle et al. (2003): Monkey; Rhesus,
30 days old*
0.5
8 h/day for 5 days, every Goblet cell metaplasia, increased AHR, and increased
5 days for a total of 11 markers of allergic asthma (e.g., eosinophilia) were
episodes observed, suggesting that episodic exposure to 03
alters postnatal morphogenesis and epithelial
differentiation and enhances the allergic effects of
house dust mite allergen in the lungs of infant
primates.
Larson etal. (2004'
Monkey;Macaca
mulatta, 30 days
old*
0.5
11 episodes of 5 days
each, 8 h/day followed
by 9 days of recovery
03 or 03 + house dust mite antigen caused changes
in density and number of airway epithelial nerves in
small conducting airways. Suggests episodic 03 alters
pattern of neural innervation in epithelial compartment
of developing lungs.
Plopper et al. (2007)
Monkey; Rhesus, 0.5
30 days old*
5 months of episodic Non-significant increases airway resistance and airway
exposure; 5 days 03 responsiveness with 03 or inhaled allergen alone.
followed by 9 days of Allergen + 03 produced additive changes in both
filtered air, 8h/day. measures.
Fanucchietal. (2006'
Monkey; male
Rhesus,30 days
old
0.5
5 months of episodic Cellular changes and significant structural changes in
exposure; 5 days 03 the distal respiratory tract in infant rhesus monkeys
followed by 9 days of exposed to 03 postnatally.
filtered air, 8h/day.
Reiser etal. (1987)
Monkey; male and 0.61
female
Cynomolgus 6-7
mo old
8 h/day for 1 year Increased lung collagen content associated with
elevated abnormal cross links that were irreversibly
deposited.
1
2
3
4
5
6
7
8
9
10
11
* sex not reported
Collectively, evidence from animal studies strongly suggests that chronic O3 exposure is
capable of damaging the distal airways and proximal alveoli, resulting in lung tissue
remodeling and leading to apparent irreversible changes. Potentially, persistent
inflammation and interstitial remodeling play an important role in the progression and
development of chronic lung disease. Further discussion of the modes of action that lead
to O3-induced morphological changes can be found in Section 5.3.7. The findings
reported in chronic animal studies offer insight into potential biological mechanisms for
the suggested association between seasonal O3 exposure and reduced lung function
development in children as observed in epidemiologic studies (see Section 7.2.3).
Discussion of mechanisms involved in lifestage susceptibility and developmental effects
can be found in Section 5.4.2.4.
12
13
14
15
16
17
18
7.2.4 Pulmonary Inflammation, Injury, and Oxidative Stress
The 2006 O3 AQCD stated that the extensive human clinical and animal toxicological
evidence, together with the limited epidemiologic evidence available, suggests a causal
role for O3 in inflammatory responses in the airways. Short-term exposure epidemiologic
studies discussed earlier in Section 6.2.3.2 show consistent associations of O3 exposure
and increased airway inflammation and oxidative stress. Further discussion of the
mechanisms underlying inflammation and oxidative stress responses can be found in
Section 5.3.3. Though the majority of recent studies focus on short-term exposures,
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1 several epidemiologic and toxicology studies of long-term exposure add to observations
2 of O 3 -induced inflammation and inj ury.
3 Inflammatory markers and peak expiratory pulmonary function were examined in 37
4 allergic children with physician-diagnosed mild persistent asthma in a highly polluted
5 urban area in Italy and then again 7 days after relocation to a rural location with
6 significantly lower pollutant levels (Renzetti et al., 2009). The authors observed a
7 fourfold decrease in nasal eosinophils and a statistically significant decrease in fractional
8 exhaled nitric oxide along with an improvement in lower airway function. Several
9 pollutants were examined, including PMi0, NO2, and O3, though pollutant-specific
10 results were not presented. These results are consistent with studies showing that traffic -
11 related exposures are associated with increased airway inflammation and reduced lung
12 function in children with asthma and contribute to the notion that this negative influence
13 may be rapidly reversible. Exhaled NO (eNO) has been shown to be a useful biomarker
14 for airway inflammation in large population-based studies (Linn et al., 2009). Thus, while
15 the time scale of 7 days between examinations for eNO needs to be evaluated for
16 appropriateness, the results suggest that inflammatory responses are reduced when O3
17 levels are decreased.
18 Chest radiographs (CXR) of 249 children in Mexico City who were chronically exposed
19 to O3 and PM25 were analyzed by Calderon-Garciduenas et al. (2006). They reported an
20 association between chronic exposures to O3 and other pollutants and a significant
21 increase in abnormal CXR's and lung CTs suggestive of a bronchiolar, peribronchiolar,
22 and/or alveolar duct inflammatory process, in clinically healthy children with no risk
23 factors for lung disease. These CXR and CT results should be viewed with caution
24 because it is difficult to attribute effects to O3 exposure.
25 In a cross-sectional study, Wood et al. (2009) examined the association of outdoor air
26 pollution with respiratory phenotype (PiZZ type) in alpha 1-Antitrypsin deficiency (a-
27 ATD) from the U.K. a-ATD registry. In total, 304 PiZZ subjects underwent full lung
28 function testing and quantitative high-resolution computed tomography to identify the
29 presence and severity of COPD - emphysema. Mean annual air pollution data for 2006
30 was matched to the location of patients' houses and used in regression models to identify
31 phenotypic associations with pollution controlling for covariates. Relative trends in O3
32 levels were assessed to validate use of a single year's data to indicate long-term exposure
33 and validation; data showed good correlations between modeled and measured data
34 (Stedman and Kent. 2008). Regression models showed that estimated higher exposure to
35 O3 exposure was associated with worse gas transfer and more severe emphysema, albeit
36 accounting for only a small proportion of the lung function variability. This suggests that
37 a gene-specific group demonstrates a long-term O3 exposure effect.
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1 The similarities of nonhuman primates to humans make them attractive models in which
2 to study the effects of O3 on the respiratory tract. The nasal mucous membranes, which
3 protect the more distal regions of the respiratory tract, are susceptible to injury from O3.
4 Carey et al. (2007) conducted a study of O3 exposure in infant rhesus macaques, whose
5 nasal airways closely resemble that of humans. Monkeys were exposed either acutely for
6 5 days (8 h/day) to 0.5 ppm O3, or episodically for several biweekly cycles alternating
7 5 days of 0.5 ppm O3 with 9 days of filtered air (0 ppm O3), designed to mimic human
8 exposure (70 days total). All monkeys acutely exposed to O3 had moderate to marked
9 necrotizing rhinitis, with focal regions of epitheliar exfoliation, numerous infiltrating
10 neutrophils, and some eosinophils. The distribution, character, and severity of lesions in
11 episodically exposed monkeys were similar to that of acutely exposed animals. Neither
12 group exhibited mucous cell metaplasia proximal to the lesions, a protective adaptation
13 observed in adult monkeys exposed continuously to 0.3 ppm O3 in another study
14 (Harkema et al.. 1987a). Adult monkeys also exhibit attenuation of inflammatory
15 responses with continued daily exposure (Harkema et al.. 1987a). but inflammation did
16 not resolve over time in young episodically exposed monkeys(Carey et al.. 2011).
17 Inflammation in conducting airways has also been observed in rats chronically exposed to
18 O3. Using an agar-based technique to fill the alveoli so that only the rat bronchi are
19 lavaged, a 90-day exposure of rats to 0.8 ppm O3 (8 h/day) elicited significantly elevated
20 pro-inflammatory eicosanoids PGE2 and 12-HETE in the conducting airway compared to
21 filtered air-exposed rats (Schmelzer et al.. 2006).
7.2.5 Allergic Responses
22 The association of air pollutants with childhood respiratory allergies was examined in the
23 U.S. using the 1999-2005 National Health Interview Survey of approximately 70,000
24 children, and ambient air pollution data from the U.S. EPA, with monitors within 20
25 miles of each child's residential block (Parker et al., 2009). The authors examined the
26 associations between the reporting of respiratory allergy or hay fever and medium-term
27 exposure to O3 over several summer months, controlling for demographic and geographic
28 factors. Increased respiratory allergy/hay fever was associated with increased O3 levels
29 (adjusted OR per 10 ppb = 1.20; [95% CI: 1.15,1.26]). These associations persisted after
30 stratification by urban-rural status, inclusion of multiple pollutants (O3, SO2, NO2, PM),
31 and definition of exposure by differing exposure radii; smaller samples within 5 miles of
32 monitors were remarkably similar to the primary results. No associations between the
33 other pollutants and the reporting of respiratory allergy/hay fever were apparent.
34 Ramadour et al. (2000) reported no relationship between O3 levels and rhinitis symptoms
35 and hay fever. Hwang et al. (2006) report the prevalence of allergic rhinitis (adjusted OR
Draft - Do Not Cite or Quote 7-25 September 2011
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1 per 10 ppb = 1.05; [95% CI: 0.98, 1.12]) in a large cross-sectional study in Taiwan. In a
2 large cross-sectional study in France, Penard-Morand et al. (2005) reported a positive
3 relationship between lifetime allergic rhinitis and O3 exposure in a two-pollutant model
4 with NO2. These studies related positive outcomes of allergic response and O3 exposure
5 but with variable strength for the effect estimates. A toxicological study reported that
6 five weeks of continuous exposure to 0.4 ppm O3 (but not 0.1 or 0.2 ppm O3) augmented
7 sneezing and nasal secretions in a guinea pig model of nasal allergy (lijima and
8 Kobayashi. 2004). Nasal eosinophils, which participate in allergic disease and
9 inflammation, and allergic antibody levels in serum were also elevated by exposure to
10 concentrations as low as 0.2 ppm (lijima and Kobayashi. 2004).
11 Nasal eosinophils were observed to decrease by fourfold in 37 atopic, mildly asthmatic
12 children 7 days after relocation from a highly polluted urban area in Italy to a rural
13 location with significantly lower pollutant levels (Renzetti et al.. 2009). Inflammatory
14 and allergic effects of O3 exposure (30 day mean) such as increased eosinophil levels
15 were observed in children in an Austrian study (Frischer et al.. 2001). Episodic exposure
16 of infant rhesus monkeys to 0.5 ppm O3 for 5 months appears to significantly increase the
17 number and proportion of eosinophils in the blood and airways (lavage) [protocol
18 described above in 7.2.3.1 for Fanucchi et al. (2006)] (Maniar-Hew et al.. 2011). These
19 changes were not evident at 1 year of age (6 months after O3 exposure ceased). Increased
20 eosinophils levels have also been observed after acute or prolonged exposures to O3 in
21 adult bonnet and rhesus monkeys (Hyde et al.. 1992; Eustis et al.. 1981).
22 Total IgE levels were related to air pollution levels in 369 adult asthmatics in five French
23 centers using generalized estimated equations (GEE) as part of the EGEA study described
24 earlier (Rage et al.. 2009b). Geostatistical models were performed on 4x4 km grids to
25 assess individual outdoor air pollution exposure that was assigned to subject's home
26 address. Ozone concentrations were positively related to total IgE levels and an increase
27 of 5 ppb of O3 resulted in an increase of 20.4% (95% CI: 3.0, 40.7) in total IgE levels.
28 Nearly 75% of the subjects were atopic. In two-pollutant models including O3 and NO2,
29 the O3 effect estimate was decreased by 25% while the NO2 effect estimate was decreased
30 by 57%. Associations were not sensitive to adjustment for covariates or the season of IgE
31 measurements. These cross-sectional results suggest that exposure to O3 may increase
32 total IgE in adult asthmatics.
33 Although very few toxicological studies of long-term exposure examining allergy are
34 available, short-term exposure studies in rodents and nonhuman primates demonstrate
35 allergic skewing of immune responses and enhanced IgE production. Due to the
36 persistent nature of these responses, the short-term toxicological evidence lends
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1 biological plausibility to the limited epidemiologic findings of an association between
2 long-term O3 exposure and allergic outcomes.
7.2.6 Host Defense
3 Short-term exposures to O3 have been shown to cause decreases in host defenses against
4 infectious lung disease in animal models. However, acute O3-induced suppression of
5 alveolar phagocytosis and immune functions observed in animals appears to be transient
6 and attenuated with continuous or repeated exposures. Chronic exposures (weeks,
7 months) of 0.1 ppm do not cause greater effects on infectivity than short exposures, due
8 to defense parameters becoming reestablished with prolonged exposures, although
9 chronic exposure has been shown to slow alveolar clearance. No detrimental effects were
10 seen with a 120-day exposure to 0.5 ppm O3 on acute lung injury from influenza virus
11 administered immediately before O3 exposure started. However, O3 was shown to
12 increase the severity of postinfluenzal alveolitis and lung parenchymal changes (Jakab
13 and Bassett. 1990). Little new evidence has become available to address the effects of
14 long-term exposure on host defense mechanisms. However, a recent study by Maniar-
15 Hew et al. (2011) demonstrated that the immune system of infant rhesus monkeys
16 episodically exposed to 0.5 ppm O3 for 5 months1 appeared to be altered in ways that
17 could diminish host defenses. Reduced numbers of circulating leukocytes were observed,
18 particularly polymorphonuclear leukocytes (PMNs) and lymphocytes, which were
19 decreased in the blood and airways (bronchoalveolar lavage). These changes did not
20 persist at 1 year of age (6 months postexposure); rather, increased numbers of monocytes
21 were observed at that time point. Challenge with LPS, a bacterial ligand that activates
22 monocytes and other innate immune cells, elicited lower responses in O3-exposed
23 animals even though the relevant reactive cell population was increased. This was
24 observed in both an in vivo inhalation challenge and an ex vivo challenge of peripheral
25 blood mononuclear cells. Thus a decreased ability to respond to pathogenic signals was
26 observed six months after O3 exposure ceased, in both the lungs and periphery.
7.2.7 Respiratory Mortality
27 A limited number of epidemiologic studies have assessed the relationship between long-
28 term exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient
29 amount of evidence existed "to suggest a causal relationship between chronic O3
30 exposure and increased risk for mortality in humans" (U.S. EPA. 2006b). Though total
1 Exposure protocol is described above in Section 7.2.3.1 for Fanucchi et al. (2006)
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1 and cardio-pulmonary mortality were considered in these studies, respiratory mortality
2 was not specifically considered. In the most recent follow-up analysis of the ACS cohort
3 (Jerrett et al.. 2009). cardiopulmonary deaths were subdivided into respiratory and
4 cardiovascular, separately, as opposed to combined in the Pope et al. (2002) work. A 10-
5 ppb increment in exposure to O3 elevated the risk of death from respiratory causes and
6 this effect was robust to the inclusion of PM2 5. The association between increased O3
7 concentrations and increased risk of death from respiratory causes was insensitive to the
8 use of a random-effects survival model allowing for spatial clustering within the
9 metropolitan area and state of residence, and to adjustment for several ecologic variables
10 considered individually. Additionally, a recent study (Zanobetti and Schwartz. In Press)
11 observed an association between long-term exposure to O3 and elevated risk of mortality
12 among Medicare enrollees that had previously experienced an emergency hospital
13 admission due to COPD.
7.2.8 Summary and Causal Determination
14 The epidemiologic studies reviewed in the 2006 O3 AQCD detected no associations
15 between long-term (annual) O3 exposures and asthma-related symptoms, asthma
16 prevalence, or allergy to common aeroallergens among children after controlling for
17 covariates. Little evidence was available to relate long-term exposure to current ambient
18 O3 concentrations to deficits in the growth rate of lung function in children. Additionally,
19 limited evidence was available evaluating the relationship between long-term O3 levels
20 and pulmonary inflammation and other endpoints. From toxicological studies, it appeared
21 that O3-induced inflammation tapered off during long-term exposures, but that
22 hyperplastic and fibrotic changes remained elevated and in some cases even worsened
23 after a postexposure period in clean air. Episodic exposures were also known to cause
24 more severe pulmonary morphologic changes than continuous exposure (U.S. EPA.
25 2006b).
26 The new epidemiologic evidence base consists of studies using a variety of designs and
27 analysis methods evaluating the relationship between long-term annual measures of
28 exposure to ambient O3 and measures of respiratory morbidity conducted by different
29 research groups in different locations. See Table 7-2 for O3 concentrations associated
30 with selected studies. The positive results from various designs and locations support an
31 association between long-term O3 concentrations and respiratory morbidity.
32 New studies examined the relationship between long-term O3 exposure and new onset
33 asthma in children. Studies have provided evidence for a relationship between different
34 genetic variants (HMOX, GST, ARG) that, in combination with O3 exposure, are related
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1
2
3
4
5
6
to new onset asthma (Islam et al., 2009; Salam et al., 2009; Islam etaL 2008). These
studies involve two separate cohorts in 12 California communities of the CHS. These
prospective cohort studies represent strong evidence because they are methodologically
rigorous epidemiology studies. The stratified analysis for the two independent fourth-
grade cohorts of the study population recruited in 1993 and 1996 yielded consistent
results and provided replication in independent groups of children.
Table 7-2 Summary of selected key new studies examining annual ozone
exposure and respiratory health effects
Study; Health Effect; Location
Akinbami et al. (2010): current asthma
United States
Hwang et al. (2005): prevalence of asthma
Taiwan
Islam et al.(2008): new-onset asthma;
CHS
Islam et al. (2009): new-onset asthma; CHS
Salam et al. (2009); childhood onset asthma; CHS
Lin et al. (2008b): first asthma hospital admission;
New York State - 10 regions
Moore et al. (2008): asthma
hospital admissions; South Coast Basin
Meng et al. (2010):
asthma ED visits or hospitalizations;
San Joaquin Valley, CA
Lee et al. (2009b):
bronchitic symptoms in asthmatic children; CHS
Rage et al. (2009b):
asthma severity; five French cities
Jacquemin et al. (In Press):
asthma control in adults; five French cities
Wentenetal. (2009):
respiratory school absence, U.S.
Annual Mean Os Concentration (ppb)
12 month median 39.8
Mean 23. 14
55.2 high vs. 38.4 low communities
10:00 a.m. to 6:00 p.m.
55.2 high vs. 38.4 low communities
10:00 a.m. to 6:00 p.m.
O3 greater than or less than 50 ppb
Range of mean O3 concentrations over the
10 New York Regions 37.51 to 47.78
Median 87.8 ppb
Median 30.3 ppb
Above and below 50 ppb
Mean 30 ppb
Median 46.9 ppb;
Median 46.9 ppb; 10:00 a.m. - 6:00 p.m.
O3 Range
(PPb)
Percent! les
IQR35.9to
43.7
Range 18.65 to
31.17
See left
See left
See left
See left
Range 28.6 to
199.9
25-75% range
27.1 to 34.0
See left
25th-75th
21-36
25th-75th
41-52
Min-Max
27.6-65.3
9
10
11
12
13
14
15
Studies using a cross-sectional design provide support for a relationship between long-
term O3 exposure and health effects in asthmatics. A long-term O3 exposure study relates
bronchitic symptoms to TNF-308 genotype asthmatic children with ambient O3 exposure
in the CHS (Lee et al., 2009b). A study in five French cities reports effects on asthma
severity related to long-term O3 exposure (Rage etal.. 2009a). A follow-up study of this
cohort (Jacquemin et al.. In Press) supports an effect of long-term O3-sum exposure on
asthma control in adulthood in subjects with pre-existing asthma. Akinbami et al. (2010)
and Hwang et al. (2005) provides further evidence relating O3 exposures and the risk of
asthma. For the respiratory health of a cohort based on the general U.S. population, risk
Draft - Do Not Cite or Quote
7-29
September 2011
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1 of respiratory-related school absences was elevated for children with the CAT and MPO
2 variant genes related to communities with high ambient O3 levels (Wenten et al.. 2009).
3 Chronic O3 exposure was related to first childhood asthma hospital admissions in a
4 positive concentration-response relationship in a New York State birth cohort (Lin et al..
5 2008b). A separate hospitalization cross-sectional study in San Joaquin Valley in
6 California reports similar findings (Meng et al., 2010). Another study relates asthma
7 hospital admissions to quarterly average O3 in the South Coast Air Basin of California
8 (Moore et al.. 2008).
9 Information from toxicological studies indicates that long term exposure to O3 during
10 gestation or development can result in irreversible morphological changes in the lung,
11 which in turn can influence pulmonary function. Studies by Plopper and colleagues have
12 demonstrated changes in pulmonary function and airway morphology in adult and infant
13 nonhuman primates repeatedly exposed to environmentally relevant concentrations of O3
14 (Tanucchi et al.. 2006: Joad et al.. 2006: Schelegle et al.. 2003: Harkema et al.. 1987b).
15 This nonhuman primate evidence of an O3-induced change in airway responsiveness
16 supports the biologic plausibility of long term exposure to O3 contributing to the adverse
17 effects of asthma in children. Results from epidemiologic studies examining long-term
18 O3 exposure and pulmonary function effects are inconclusive with some new studies
19 relating effects at higher exposure levels. The results from the CHS described in the 2006
20 O3 AQCD remain the definitive line of evidence. Other cross-sectional studies provide
21 mixed results.
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1
2
3
4
5
6
Table 7-3 Studies providing evidence concerning potential confounding by
PM for available endpoints
Study
Exposure
Endpomt
Hwang et al. (2005)
10ppbO3
Asthma risk in children
Jacquemin et al. (In
presS) IQR 25-38 ppb O3
, , • , „ summer
Asthma control in adults
Lin et al. (2008b)
Asthma admissions in IQR 2.5%
children
Akinbami et al. (2010)
Asthma prevalence in IQR 35.9-43.7 ppb
children
Lee et al. (2009b)
Bronchitic symptoms High O3 >50 ppb
asthmatics
Rage et al. (2009a)
Asthma severity in IQR 28.5-33.9 ppb
adults
Meng et al. (2007)
1 ppm
Asthma control
Meng et al. (201 0)
Asthma ED visits, 10 ppb
Hospitalization
Karretal. (2007)
Bronchiolitis 10 ppb
Hospitalization
Rojas-Martinez et al.
(2007)
i-i-w , i % ™ r -t 11. 3 ppb IQR
FEN/! (mL) Deficit ^
Girls
Parker et al. (2009)
10 ppb
Respiratory allergy
Single
Pollutant O3
1.138
(1.001, 1.293
1.69
(1 .22, 2.34)
1.16
(1.15, 1.17)
1.56
(1.15,2.10)
1.42
(0.75, 2.70)
2.53
(1 .69, 3.79)
1.70
(0.91,3.18)
1.49
(1.05,2.11)
0.92
(0.88, 0.96)
-24
(-30, -19)
1.24
(1.15, 1.34)
Single
Pollutant PM
0.934
(0.909, 0.960)
1.33
(1 .06, 1 .67)
NA
PM2.5
1.43
(0.98,2.10)
NA
NA
PM10 2.06
(1.17,3.61)
women
PM10
1.29
(0.99, 1 .69)
1.09
(1.04, 1.14)
PM10
IQR
36.4 ug/m3
-29(-36, -21)
1.23
(1 .04, 1 .46)
O3 with PM
PM10
1.253(1.089,1.442)
PM10
1.50(1.07,2.11)
Air Quality Index
1.24(1.23, 1.25)
Adjusted for
S02,PM2.5,PM10
1.86(1.02-3.40)
Adjusted for PM25,
PM10
1.36(0.91-2.02)
No substantial
differences
PM10, PM2.5
No PM data
Three pollutant (O3,
NO2, SO2)
2.74(1.68,4.48)
Did not differ
Did not differ
PM2.5
1.02(0.94, 1.10)
-17 (-23, -12)
Multi-pollutant
1.18(1.09,1.27)
PM with O3
0.925
(0.899, 0.952)
1.28
(1 .06, 1 .55)
NA
PM2.5
1.24(0.70-2.21)
PM2.5
1.26(0.80-1.98)
NA
NA
NA
NA
1.09
(1.03, 1.15)
-24 (-31 ,
-16)
1.29(1.07,1.56)
The highest quartile is shown for all results.
NA= not available
Several studies (see Table 7-3) provide results from studies that adjusted for potential
confounders, presenting results for both O3 and PM (single and multipollutant models) as
well as other pollutants where PM effects were not provided. As shown in the table, O3
associations are generally robust to adjustment for potential confounding by PM.
The 2006 O3 AQCD concluded that the extensive human clinical and animal
toxicological evidence, together with the limited epidemiologic evidence available,
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7-31
September 2011
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1 suggests a causal role for short-term O3 exposure in inflammatory responses in the
2 airways. Though the majority of recent studies focus on short-term exposures, several
3 epidemiologic and toxicological studies of long-term exposure add to observations of O3 -
4 induced inflammation and injury. Toxicological studies in rodents and nonhuman
5 primates indicate that chronic O3 exposure causes structural changes in the respiratory
6 tract, and simulated seasonal exposure studies suggest that such exposures might have
7 cumulative impacts. The strongest epidemiologic evidence for a relationship between
8 long-term O3 exposure and respiratory morbidity is provided by new studies that
9 demonstrate associations between long-term measures of O3 exposure and new-onset
10 asthma in children and increased respiratory symptom effects in asthmatics. While there
11 are currently a limited number of studies in this data base, the U.S. multi-community
12 prospective cohort studies are methodologically rigorous epidemiologic studies. Asthma
13 risk is related to complex relationships between genetic variability, environmental O3
14 exposure, and behavior. The genes, evaluated in these studies, are both key candidates in
15 the oxidative stress pathway and have been shown to play an important role in asthma
16 development. Reduced risk for asthma development is reported in some studies in
17 children living in low- O3 communities. Mean O3 concentrations in the studies (10:00
18 a.m. to 6:00 p.m.) ranged from 28.6 to 45.5 ppb in low O3 communities
19 (mean = 38.4 ppb) and from 46.5 to 64.9 ppb in high O3 communities (mean = 55.2 ppb).
20 These CHS multi-community studies form a foundation for the evidence base in which
21 findings for several genes indicate the breath of the evidence across different gene
22 variants. The several other studies with different designs, analysis, locations and
23 researchers provide a cumulative collective body of evidence informing these
24 relationships. The other studies in the new data base provide coherent evidence for long-
25 term O3 exposure and respiratory morbidity effects such as first asthma hospitalization
26 and respiratory symptoms in asthmatics. Studies considering other pollutants provide data
27 suggesting that the effects related to O3 are independent from potential effects of the
28 other pollutants. Some studies provide evidence for a positive concentration-response
29 relationship. Short-term studies provide supportive evidence with increases in respiratory
30 symptoms and asthma medication use, hospital admissions and ED visits for all
31 respiratory outcomes and asthma, and decrements in lung function in children. The above
32 discussion of the recent epidemiologic and toxicological data base provides a compelling
33 case to support the hypothesis that a relationship exists between long-term exposure to
34 ambient O3 and measures of respiratory morbidity. The 2006 O3 AQCD concluded the
35 evidence was suggestive but inconclusive at that time. The new epidemiological data
36 base, combined with toxicological studies in rodents and nonhuman primates,
37 provides biologically plausible evidence that there is likely to be causal
38 relationship between long-term exposure to O3 and respiratory morbidity.
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7.3 Cardiovascular Effects
7.3.1 Cardiovascular Disease
7.3.1.1 Cardiovascular Epidemiology
1 Long-term exposure to O3 and its effects on cardiovascular morbidity were not
2 considered in the 2006 O3 AQCD. However, recent studies have assessed the chronic
3 effects of O3 exposure on cardiovascular morbidity (Chuang etal.. 2011; Forbes et al.
4 2009a; Chen etal.. 2007). The association between O3 exposure and markers of lipid
5 peroxidation and antioxidant capacity was examined among 120 nonsmoking healthy
6 college students, aged 18-22 years, from the University of California, Berkeley (Feb-Jun
7 2002) (Chen et al.. 2007). By design, students were chosen from geographic areas so they
8 had experienced different levels of O3 over their lifetimes and during recent summer
9 vacation in either greater Los Angeles (LA) or the San Francisco Bay Area (SF). A
10 marker of lipid peroxidation, 8-isoprostane (8-iso-PGF) in plasma, was assessed. This
11 marker is formed continuously under normal physiological conditions but has been found
12 at elevated concentrations in response to environmental exposures. A marker of overall
13 antioxidant capacity, ferric reducing ability of plasma (FRAP), was also measured. The
14 lifetime O3 exposure estimates (estimated monthly average) did not show much overlap
15 between the two geographic areas [median (range): LA, 42.9 ppb (28.5-65.3); SF, 26.9
16 ppb (17.6-33.5)]. Estimated lifetime O3 exposure was related to 8-iso-PGF [(3 = 0.025
17 (pg/mL)/8-h ppb O3, p = 0.0007]. For the 17-ppb cumulative lifetime O3 exposure
18 difference between LA and SF participants, there was a 17.41-pg/mL (95% CI: 15.43,
19 19.39) increase in 8-iso-PGF. No evidence of association was observed between lifetime
20 O3 exposure and FRAP [(3 = -2.21 (pg/mL)/8-h ppb O3, p = 0.45]. The authors note that
21 O3 was highly correlated with PM10_2 5 and NO2 in this study population; however, their
22 inclusion in the O3 models did not substantially modify the magnitude of the associations
23 with O3. Because the lifetime exposure results were supported by shorter-term exposure
24 results from analyses considering O3 concentrations up to 30 days prior to sampling, the
25 authors conclude that persistent exposure to O3 can lead to sustained oxidative stress and
26 increased lipid peroxidation. However, because there was not much overlap in lifetime
27 O3 exposure estimates between LA and SF, it is possible that the risk estimates involving
28 the lifetime O3 exposures could be confounded by unmeasured factors related to other
29 differences between the two cities.
30 Forbes et al. (2009a) used the annual average exposures to assess the relationship
31 between chronic ambient air pollution and levels of fibrinogen and C-reactive protein
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1 (CRP) in a cross-sectional study conducted in England. Data were collected from the
2 Health Survey of England for 1994, 1998, and 2003. The sampling strategy was designed
3 to obtain a representative sample of the English population; however, due to small group
4 sizes, only data from white ethnic groups were analyzed. For analyses, the annual
5 concentrations of O3 were averaged for the year of data collection and the previous year
6 with the exception of 1994 (because pollutant data were not available for 1993). Median
7 O3 concentrations were 26.7 ppb, 25.4 ppb, and 28 ppb for 1994, 1998, and 2003,
8 respectively. Year specific adjusted effect estimates were created and combined in a
9 meta-analysis. No evidence of association was observed for O3 and levels of fibrinogen
10 or CRP (e.g., the combined estimates for the percent change in fibrinogen and CRP for a
11 10 ppb increase in O3 were -0.28 [95% CI: -2.43, 1.92] and -3.05 [95% CI: -16.10,
12 12.02], respectively).
13 A study was performed in Taiwan to examine the association between long-term O3
14 concentrations and blood pressure and blood markers using the Social Environment and
15 Biomarkers of Aging Study (SEBAS) (Chuang et al.. 2011). Individuals included in the
16 study were 54 years of age and older. The mean annual O3 concentration during the study
17 period was 22.95 ppb (SD 6.76 ppb). Positive associations were observed between O3
18 concentrations and both systolic and diastolic blood pressure [changes in systolic and
19 diastolic blood pressure were 21.51mmHg (95% CI: 16.90, 26.13) and 20.56 mmHg
20 (95% CI: 18.14, 22.97) per 8.95 ppb increase in O3, respectively). Increased O3
21 concentrations were also associated with increased levels of total cholesterol, fasting
22 glucose, hemoglobin Ale, and neutrophils. No associations were observed between O3
23 concentrations and triglyceride and IL-6 levels. The observed associations were reduced
24 when other pollutants were added to the models. Further research will be important for
25 understanding the effects, if any, of chronic O3 exposure on cardiovascular morbidity
26 risk.
7.3.1.2 Cardiovascular Toxicology
27 Three new studies have investigated the cardiovascular effects of long-term exposure to
28 O3 in animal models (See Table 7-4 for study details). In addition to the short-term
29 exposure effects described in Section 6.3.3, a recent study found that O3 exposure in
30 genetically hyperlipidemic mice enhanced aortic atherosclerotic lesion area compared to
31 air exposed controls (Chuang et al.. 2009). Chuang et al. (2009) not only provided
32 evidence for increased atherogenesis in susceptible mice, but also reported an elevated
33 vascular inflammatory and redox state in wild-type mice and infant primates
34 (Section 6.3.3.2). This study is compelling in that it identifies biochemical and cellular
3 5 events responsible for transducing the airway epithelial reactions of O3 into
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1 proinflammatory responses that are apparent in the extrapulmonary vasculature (Cole and
2 Freeman. 2009).
3 Another recent study provides further evidence for increased vascular inflammation and
4 oxidation and long term effects in the extrapulmonary space. Rats episodically exposed to
5 O3 for 16 weeks presented marked increases in gene expression of biomarkers of
6 oxidative stress, thrombosis, vasoconstriction, and proteolysis (Kodavanti et al., 2011).
7 Ozone exposure upregulated aortic mRNA expression of heme oxygenase-1 (HO-1),
8 tissue plasminogen activator (tPA), plasminogen activator inhibitor-1 (PAI-1), von
9 Willebrand factor (vWf), thrombomodulin, endothelial nitric oxide synthase (eNOS),
10 endothelin-1 (ET-1), matrix metalloprotease-2 (MMP-2), matrix metalloprotease-3
11 (MMP-3), and tissue inhibitor of matrix metalloprotease-2 (TIMP-2). In addition, O3
12 exposure depleted some cardiac mitochondrial phospholipid fatty acids (C16:0 and
13 CIS: 1), which may be the result of oxidative modifications. The authors speculate that
14 oxidatively modified lipids and proteins produced in the lung and heart promote vascular
15 pathology through activation of lectin-like oxidized-low density lipoprotein receptor-1
16 (LOX-1). Activated LOX-1 induces expression of a number of the biomarkers induced by
17 O3 exposure and is considered pro-atherogenic. Both LOX-1 mRNA and protein were
18 increased in mouse aorta after O3 exposure. This study provides a possible pathway and
19 further support to the observed O3 induced atherosclerosis.
20 Vascular occlusion resulting from atherosclerosis can block blood flow through vessels
21 causing ischemia. The restoration of blood flow or reperfusion can cause injury to the
22 tissue from subsequent inflammation and oxidative damage. Ozone exposure enhanced
23 the sensitivity to myocardial ischemia-reperfusion (I/R) injury in rats while increasing
24 oxidative stress levels and pro-inflammatory mediators and decreasing production of anti-
25 inflammatory proteins (Perepu et al.. 2010). Both long- and short-term O3 exposure
26 decreased the left ventricular developed pressure, rate of change of pressure
27 development, and rate of change of pressure decay and increased left ventricular end
28 diastolic pressure in isolated perfused hearts (Section 6.3.3.2 for short-term exposure
29 discussion). In this ex vivo heart model, O3 induced oxidative stress by decreasing SOD
30 enzyme activity and increasing malondialdehyde levels. Ozone also elicited a
31 proinflammatory state evident by an increase in TNF-a and a decrease in the anti-
32 inflammatory cytokine IL-10. The authors conclude that O3 exposure will result in a
33 greater I/R injury.
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Table 7-4
Study
Chuang et al. (2009)
Kodavanti et al.
(2011)
Perepu et al. (201 0)
Characterization of study details for Section 7.3.1.2
Model
Mice; ApoE-/-; M;
6 weeks
Rat; Wistar; M;
10-1 2 weeks
Rat; Sprague-Dawley;
Weight: 50-75 g
_ . . Exposure
°3 Duration
n ,- 8 wks, 5 days/week,
u'° 8 h/day
n . 16 wks, 1 day/week,
u'4 5 h/day
0.8 56 days, 8 h/day
Effects
Enhanced aortic atherosclerotic lesion
area compared to air controls.
Increased vascular inflammation and
oxidative stress, possibly through
activation of LOX-1 signaling.
Enhanced the sensitivity to myocardial
I/R injury while increasing oxidative
stress and pro-inflammatory mediators
and decreasing production of anti-
inflammatory proteins.
No previous studies investigated cardiovascular effects from long-term exposure to O3.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
7.3.2 Cardiac Mortality
A limited number of epidemiologic studies have assessed the relationship between long-
term exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient
amount of evidence existed "to suggest a causal relationship between chronic O3
exposure and increased risk for mortality in humans" (U.S. EPA. 2006b). Though total
and cardio-pulmonary mortality were considered in these studies, cardiovascular
mortality was not specifically considered. In the most recent follow-up analysis of the
ACS cohort (Jerrett et al.. 2009). cardiopulmonary deaths were subdivided into
respiratory and cardiovascular, separately, as opposed to combined in the Pope et al.
(2002) work. A 10-ppb increment in exposure to O3 elevated the risk of death from the
cardiopulmonary, cardiovascular, and ischemic heart disease. Inclusion of PM25 as a
copollutant attenuated the association with exposure to O3 for all of the cardiovascular
endpoints to become null. Additionally, a recent study (Zanobetti and Schwartz. In Press)
observed an association between long-term exposure to O3 and elevated risk of mortality
among Medicare enrollees that had previously experienced an emergency hospital
admission due to congestive heart failure (CHF) or myocardial infarction (MI).
16
17
18
19
20
21
7.3.3 Summary and Causal Determination
Previous AQCDs did not address the cardiovascular effects of long-term O3 exposure due
to limited data availability. The evidence remains limited; however the emerging data is
supportive of a role for O3 in chronic cardiovascular diseases. Few epidemiologic studies
have investigated cardiovascular morbidity after long-term O3 exposure, and the majority
only assessed cardiovascular disease related biomarkers. A study on O3 and
cardiovascular mortality reported no association after adjustment for PM2 5 levels.
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1 Further epidemiologic studies on cardiovascular morbidity and mortality after long-term
2 exposure have not been published.
3 lexicological evidence on long-term O3 exposure is also limited but three strong
4 toxicological studies have been published since the previous AQCD. These studies
5 provide evidence for O3 enhanced atherosclerosis and I/R injury, corresponding with
6 development of a systemic oxidative, proinflammatory environment. Further discussion
7 of the mechanisms that may lead to cardiovascular effects can be found in Section 5.3.8.
8 Although questions exist for how O3 inhalation causes systemic effects, a recent study
9 proposes a mechanism for development of vascular pathology that involves activation of
10 LOX-1 by O3 oxidized lipids and proteins. This activation may also be responsible for O3
11 induced changes in genes involved in proteolysis, thrombosis, and vasoconstriction.
12 Taking into consideration the findings of toxicological studies, and the emerging
13 evidence from epidemiologic studies, the generally limited body of evidence is
14 suggestive of a causal relationship between long-term exposures to O3 and
15 cardiovascular effects.
7.4 Reproductive and Developmental Effects
16 Although the body of literature is growing, the research focusing on adverse birth
17 outcomes is small. Among these studies, various measures of birth weight and fetal
18 growth, such as low birth weight (LEW), small for gestational age (SGA), and
19 intrauterine growth restriction (IUGR), and preterm birth (<37-week gestation; [PTB])
20 have received more attention in air pollution research, while congenital malformations
21 are less studied. There are also new studies on reproductive and developmental effects.
22 Infants and fetal development processes may be particularly susceptible to O3-induced
23 health effects, and although the physical mechanisms are not fully understood, several
24 hypotheses have been proposed; these include: oxidative stress, systemic inflammation,
25 vascular dysfunction and impaired immune function (Section 5.3). Study of these
26 outcomes can be difficult given the need for detailed exposure data and potential
27 residential movement of mothers during pregnancy. Air pollution epidemiologic studies
28 reviewed in the 2006 O3 AQCD examined impacts on birth-related endpoints, including
29 intrauterine, perinatal, postneonatal, and infant deaths; premature births; intrauterine
30 growth retardation; very low birth weight (weight < 1,5 00 grams) and low birth weight
31 (weight <2,500 grams); and birth defects. However, in the limited number of studies that
32 investigated O3, no associations were found between O3 and birth outcomes, with the
33 possible exception of birth defects.
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1 Several recent articles have reviewed methodological issues relating to the study of
2 outdoor air pollution and adverse birth outcomes (Chen et al.. 2010a: Woodruff etal..
3 2009; Ritz and Wilhelm. 2008; Slamaet al.. 2008). Some of the key challenges to
4 interpretation of these study results include the difficulty in assessing exposure as most
5 studies use existing monitoring networks to estimate individual exposure to ambient air
6 pollution; the inability to control for potential confounders such as other risk factors that
7 affect birth outcomes (e.g., smoking); evaluating the exposure window (e.g., trimester) of
8 importance; and limited evidence on the physiological mechanism of these effects (Ritz
9 and Wilhelm. 2008; Slama et al.. 2008). Recently, an international collaboration was
10 formed to better understand the relationships between air pollution and adverse birth
11 outcomes and to examine some of these methodological issues through standardized
12 parallel analyses in datasets from different countries (Woodruff et al.. 2010). Initial
13 results from this collaboration have examined PM and birth weight (Parker et al., 2011);
14 work on O3 has not yet been performed. Although early animal studies (Kavlock et al..
15 1980) found that exposure to O3 in the late gestation of pregnancy in rats led to some
16 abnormal reproductive performances for neonates, to date human studies have reported
17 inconsistent results for the association of ambient O3 on birth outcomes.
7.4.1 Effects on Sperm
18 A limited amount of research has been conducted to examine the association between air
19 pollution and male reproductive outcomes, specifically semen quality. To date, the
20 epidemiologic studies have considered various exposure durations before semen
21 collection that encompass either the entire period of spermatogenesis (i.e., 90 days) or
22 key periods of sperm development that correspond to epididymal storage, development of
23 sperm motility, and spermatogenesis. In an analysis conducted as part of the Teplice
24 Program, 18-year-old men residing in the heavily polluted district of Teplice in the Czech
25 Republic were found to be at greater risk of having abnormalities in sperm morphology
26 and chromatin integrity than men of similar age residing in Prachatice, a less polluted
27 district (Selevan et al.. 2000; Sram etal.. 1999). A follow-up longitudinal study
28 conducted on a subset of the same men from Teplice revealed associations between total
29 episodic air pollution and abnormalities in sperm chromatin (Rubes et al.. 2005). A
30 limitation of these studies is that they did not identify specific pollutants and their
31 concentrations.
32 More recent epidemiologic studies conducted in the U.S. have also reported associations
33 between ambient air pollution and sperm quality for individual air pollutants, including
34 O3 and PM25. In a repeated measures study in Los Angeles, CA, Sokol et al. (2006)
35 reported a reduction in average sperm concentration during three exposure windows (0-9,
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1 10-14, and 70-90 days before semen collection) associated with high ambient levels of
2 O3 in healthy sperm donors. This effect persisted under a joint additive model for O3,
3 CO, NO2 and PM10. The authors did not detect a reduction in sperm count. Hansen et al.
4 (2010) investigated the effect of exposure to O3 and PM2 5 (using the same exposure
5 windows used by Sokol et al. (2006) on sperm quality in three southeastern counties
6 (Wake County, NC; Shelby County, TN; Galveston County, TX). Outcomes included
7 sperm concentration and count, morphology, DNA integrity and chromatin maturity.
8 Overall, the authors found both protective and adverse effects, although some results
9 suggested adverse effects on sperm concentration, count and morphology.
10 The biological mechanisms linking ambient air pollution to decreased sperm quality have
11 yet to be determined, though O3-induced oxidative stress, inflammatory reactions, and
12 the induction of the formation of circulating toxic species have been suggested as
13 possible mechanisms (see Section 5. 3.8). Decremental effects on testicular morphology
14 have been demonstrated in toxicological studies with histological evidence of O3-induced
15 depletion of germ cells in testicular tissue and decreased seminiferous tubule epithelial
16 layer. Jedlinska-Krakowska et al. (2006) demonstrated histopathological evidence of
17 impaired spermatogenesis (round spermatids/ spermatocytes, giant spermatid cells, and
18 focal epithelial desquamation with denudation to the basement membrane). The exposure
19 protocol used five month old adult rats exposed to O3 as adults (0.5 ppm, 5 h/day for
20 50 days). This degeneration could be rescued by vitamin E administration, indicating an
21 antioxidant effect. Vitamin C administration had no effect at low doses of ascorbic acid
22 and exacerbated the O3-dependent damage at high doses, as would be expected as
23 vitamin C can be a radical generator instead of an antioxidant at higher doses. In
24 summary, this study provided toxicological evidence of impaired spermatogenesis with
25 O3 exposure that was rescued with certain antioxidant supplementation.
26 Overall, there is limited epidemiologic evidence for an association with O3 concentration
27 and decreased sperm concentration. A recent toxicological study provides limited
28 evidence for a possible biological mechanism (histopathology showing impaired
29 spermatogenesis) for such an association.
7.4.2 Effects on Reproduction
30 Evidence suggests that exposure to air pollutants during pregnancy is associated with
31 adverse birth outcomes, which has been attributed to the increased susceptibility of the
32 fetus due to physiologic immaturity. Gametes (i.e., ova and sperm) may be even more
33 susceptible, especially outside of the human body, as occurs with assisted reproduction.
34 Smokers require twice the number of in vitro fertilization (IVF) attempts to conceive as
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1 non-smokers (Feichtinger et al., 1997). suggesting that a preconception exposure can be
2 harmful to pregnancy. A recent study used an established national-scale, log-normal
3 kriging method to spatially estimate daily mean concentrations of criteria pollutants at
4 addresses of women undergoing their first IVF cycle and at their IVF labs from 2000 to
5 2007 in the northeastern U.S. (Legro et al., 2010). Increasing O3 concentration at the
6 patient's address was significantly associated with an increased chance of live birth
7 during ovulation induction (OR=1.13, [95% CI: 1.05, 1.22] per 10 ppb increase), but with
8 decreased odds of live birth when exposed from embryo transfer to live birth (OR=0.79,
9 [95% CI: 0.69, 0.90] per 10 ppb increase). After controlling for NO2 in a copollutant
10 model, however, O3 was no longer significantly associated with IVF failure. The results
11 of this study suggest that exposure to O3 during ovulation was beneficial (perhaps due to
12 early conditioning to O3), whereas later exposure to O3 (e.g., during gestation) was
13 detrimental, and reduced the likelihood of a live birth.
14 In most toxicological studies, reproductive success appears to be unaffected by O3
15 exposure. Nonetheless, one study has reported that 25% of the BALB/c mouse dams in
16 the highest O3 exposure group (1.2 ppm, GD9-18) did not complete a successful
17 pregnancy, a significant reduction (Sharkhuu et al.. 2011). Ozone administration
18 (continuous 0.4, 0.8 or 1.2 ppm O3) to CD-I mouse dams during the majority of
19 pregnancy (PD7-17, which excludes the pre-implantation period), led to no adverse
20 effects on reproductive success (proportion of successful pregnancies, litter size, sex
21 ratio, frequency of still birth, or neonatal mortality) (Bignami et al.. 1994). There was a
22 nearly statistically significant increase in pregnancy duration (0.8 and 1.2 ppm O3).
23 Initially, dam body weight (0.8 and 1.2 ppm), water consumption (0.4, 0.8 and 1.2 ppm
24 O3) and food consumption (0.4, 0.8 and 1.2 ppm) during pregnancy were decreased with
25 O3 exposure but these deficits dissipated a week or two after the initial exposure
26 (Bignami et al.. 1994). The anorexigenic effect of O3 exposure on the pregnant dam
27 appears to dissipate with time; the dams seem to adapt to the O3 exposure. In males, data
28 exist showing morphological evidence of altered spermatogenesis in O3 exposed animals
29 (Jedlinska-Krakowska et al.. 2006). Some evidence suggests that O3 may affect
30 reproductive success when combined with other chemicals. Kavlock et al. (1979) showed
31 that O3 acted synergistically with sodium salicylate to increase the rate of pup resorptions
32 after midgestational exposure (1.0 ppm O3, GD9-12). At low doses of O3 exposure,
33 toxicological studies show reproductive effects to include a transient anorexigenic effect
34 of O3 on gestational weight gain, and a synergistic effect of O3 on salicylate-induced pup
35 resorptions; other fecundity, pregnancy and gestation related outcomes appear unaffected
36 by O3 exposure.
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1 Collectively, there is very little epidemiologic evidence for the effect of O3 on
2 reproductive success, and the reproductive success in rats appears to be unaffected in
3 toxicological studies of O3 exposure.
7.4.3 Birth Weight
4 With birth weight routinely collected in vital statistics and being a powerful predictor of
5 infant mortality, it is the most studied outcome within air pollution-birth outcome
6 research. Air pollution researchers have analyzed birth weight as a continuous variable
7 and/or as a dichotomized variable in the form of LEW (<2,500 g [5 Ibs, 8 oz]).
8 Birth weight is primarily determined by gestational age and intrauterine growth, but also
9 depends on maternal, placental and fetal factors as well as on environmental influences.
10 In both developed and developing countries, LEW is the most important predictor for
11 neonatal mortality and is a significant determinant of postneonatal mortality and
12 morbidity. Studies report that infants who are smallest at birth have a higher incidence of
13 diseases and disabilities, which continue into adulthood (Hack and Fanaroff. 1999).
14 The strongest evidence for an effect of O3 on birth weight comes from the Children's
15 Health Study conducted in southern California. In this study, Salam et al. (2005) report
16 that maternal exposure to 24-h avg O3 concentrations averaged over the entire pregnancy
17 was associated with reduced birth weight (39.3 g decrease [95% CI: -55.8, -22.8] in birth
18 weight per 10 ppb and 8-h avg (19.2-g decrease [95% CI: -27.7, -10.7] in birth weight per
19 10 ppb). This effect was stronger for concentrations averaged over the second and third
20 trimesters. PM10, NO2 and CO concentrations averaged over the entire pregnancy were
21 not statistically significantly associated with birth weight, although CO concentrations in
22 the first trimester and PM10 concentrations in the third trimester were associated with a
23 decrease in birth weight. Additionally, the authors observed a concentration-response
24 relationship of birth weight with 24-h avg O3 concentrations averaged over the entire
25 pregnancy that was clearest above the 30-ppb level (see Figure 7-4). Relative to the
26 lowest decile of 24-h avg O3, estimates for the next 5 lowest deciles were approximately
27 -40 g to -50 g, with no clear trend and with 95% confidence bounds that included zero.
28 The highest four deciles of O3 exposure showed an approximately linear decrease in birth
29 weight, and all four 95% CIs excluded zero, and ranged from mean decreases of
30 74 grams to decreases of 148 grams.
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50
0
-50
'5 -100
-150
-200
-250
c
3
C
(
3 f
> c
_,
)
c
•)
c
)
[
)
0
20 30
24-hr 03{ppb)
40
50
Source: Salam et al. (2005)
Deficits are plotted against the decile-group-specific median O3 exposure. Error bars represent 95% CIs. Indicator variables for
each decile of O3 exposure (except the least-exposed group) were included in a mixed model.
Figure 7-4 Birthweight deficit by decile of 24-h avg O3 concentration averaged
over the entire pregnancy compared with the decile group with the
lowest Oz exposure.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
Several additional studies conducted in the U.S. and Canada also investigated the
association between ambient O3 concentrations and birth weight and report some weak
evidence for an association. Morello-Frosch et al. (2010) estimated ambient O3
concentrations throughout pregnancy and for each trimester in the neighborhoods of
women who delivered term singleton births between 1996 and 2006 in California. A 10-
ppb increase in O3 averaged across the entire pregnancy was associated with a 5.7-g
decrease (95% CI: -6.6, -4.9) in birth weight when exposures were calculated using
monitors within 10 km of the maternal address at date of birth. When the distance from
the monitor was restricted to 3 km, the decrease in birth weight associated with a 10-ppb
increase in O3 increased to 8.9 g (95% CI: -10.6, -7.1). These results persisted in
copollutant models and in models that stratified by trimester of exposure, SES, and race.
Darrow et al. (2011 a) did not observe an association with birth weight and O3
concentrations during two exposure periods of interest (i.e., the first month and last
trimester), but did find an association with reduced birth weight when examining the
cumulative air pollution concentration during the entire pregnancy period. Additionally,
they observed effect modification by race and ethnicity, such that associations between
birth weight and third-trimester O3 concentrations were significantly stronger in
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September 2011
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1 Hispanics and non-Hispanic African Americans than in non-Hispanic whites. Chen et al.
2 (2002) used 8-h avg O3 concentrations to create exposure variables based on average
3 maternal exposure for each trimester. Ozone was not found to be related to birth weight
4 in single-pollutant models, though the O3 effect during the third trimester was borderline
5 statistically significant in a copollutant model with PM10.
6 Several studies found no association between ambient O3 concentrations and birth
7 weight. Wilhelm and Ritz (2005) extended previous analyses of term LEW (Ritz et al..
8 2000; Ritz and Yu. 1999) to include the period 1994-2000. The authors examined varying
9 residential distances from monitoring stations to see if the distance affected risk
10 estimation, exploring the possibility that effect attenuation may result from local pollutant
11 heterogeneity inadequately captured by ambient monitors. As in their previous studies,
12 the authors observed associations between elevated concentrations of CO and PM10 both
13 early and late in pregnancy and risk of term LEW. After adjusting for CO and/or PMi0
14 the authors did not observe associations between O3 and term LEW in any of their
15 models. Brauer et al. (2008) evaluated the impacts of air pollution (CO, NO2, NO, O3,
16 SO2, PM2 5, PM10) on birth weight for the period 1999-2002 using spatiotemporal
17 residential exposure metrics by month of pregnancy in Vancouver, BC. Quantitative
18 results were not presented for the association between O3 and LEW, though the authors
19 observed associations that were largely protective. Dugandzic et al. (2006) examined the
20 association between LEW and ambient levels of air pollutants by trimester of exposure
21 among a cohort of term singleton births from 1988-2000. Though there was some
22 indication of an association with SO2 and PMi0, there were no effects for O3.
23 Similarly, studies conducted in Australia, Latin America, and Asia report limited
24 evidence for an association between ambient O3 and measures of birth weight. In Sydney,
25 Australia, Mannes et al. (2005) found that O3 concentrations in the second trimester of
26 pregnancy had small adverse effects on birth weight (7.5-g decrease; [95 % CI: -13.8,
27 1.2] per 10 ppb), although this effect disappeared when the analysis was limited to births
28 with a maternal address within 5 km of a monitoring station (87.7-g increase; [95% CI:
29 10.5, 164.9] per 10 ppb). Hansen et al. (2007) reported that trimester and monthly
30 specific exposures to all pollutants were not statistically significantly associated with a
31 reduction in birth weight in Brisbane, Australia. In Sao Paulo, Brazil, Gouveia et al.
32 (2004) found that O3 exhibited a small inverse relation with birth weight over the third
33 trimester (6.0-g decrease; [95% CI: -30.8, 18.8] per 10 ppb). Lin et al. (2004b) reported a
34 positive, though not statistically significant, exposure-response relationship for O3 during
35 the entire pregnancy in a Taiwanese study. In a study performed in Korea, Ha et al.
36 (2001) reported no O3 effect during the first trimester of pregnancy, but they found that
37 during the third trimester of pregnancy O3 was associated with LEW (RR=1.05 [95% CI:
38 1.02, 1.08] per 10 ppb).
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1
2
3
4
Table 7-5
Study
Salametal. (2005)
Morello-Froschetal.
(2010)
Darrow et al. (2011 a)
Chen etal. (2002)
Wilhelm and Ritz (2005)
Braueretal. (2008)
Dugandzicetal. (2006)
Mannesetal. (2005)
Hansen et al. (2007)
Gouveiaetal. (2004)
Lin et al. (2004b)
Ha et al. (2001)
Brief summary of epidemiologic studies of birth weight
Location
Sample Size
California, U.S.
(n=3,901)
California, U.S.
(n=3,545,177)
Atlanta, GA
(N=406,627)
Northern Nevada, US
(n=36,305)
Los Angeles County, CA
(n=136,134)
Vancouver, BC, Canada
(n=70,249)
Nova Scotia, Canada
(n=74,284)
Sydney, Australia
(n=138,056)
Brisbane, Australia
(n=26,617)
Sao Paulo, Brazil
(n=1 79,460)
Kaohsiung and Taipei,
Taiwan
(n=92,288)
Seoul, Korea
(n=276,763)
Mean Os (ppb)
24-h avg:
27.3
8h:
50.6
24-h avg:
23.5
8-h max:
44.8
8-h:
27.2
1-h:
21.1-22.2
24-h avg:
14
24-h avg:
21
1-h max:
31.6
8 h max:
26.7
1-h max:
31.5
24-h avg: 15.86-
47.78
8-h avg: 22.4-23.3°
Exposure assessment
ZIP code level
Nearest Monitor
(within 10,5,3km)
Population-weighted spatial
average
County level
Varying distances from monitor
Nearest Monitor
(within 10km)
Inverse Distance Weighting (IDW)
Nearest Monitor
(within 25 km)
City-wide avg and
<5 km from monitor
City-wide avg
City-wide avg
Nearest monitor
(within 3 km)
City-wide avg
Effect Estimate3
(95% Cl)
Entire pregnancy:
-39.3 g (-55.8, -22.8)
T1 : -6.1 g (-16.8, 4.8)
T2: -20.0 g (-31 .7, -8.4)
T3: -20.7 g (-32.1, -9.3)
Entire pregnancy: -5.7 g (-6.6, -
4.9)
T1:-2.1g(-2.9, -1.4)
T2: -2.3 g (-3.1, -1.5)
T3:-1.3g(-2.1,-0.6)
Entire pregnancy:
-1 2.3 g (-17.8, -6.8)
First 28 days:
-0.5 g (-3.0, 2.1)
T3: -0.9g (-4.5, 2.8)
Entire pregnancy:
20.9 g (6.3, 35.5)
T1: 23.4 g (-35.6, 82.4)
T2:-1 9.4 g (-77.0, 38.2)
T3: 7.7 g (-50.9, 66.3)
T1:NR
T3:NR
6 weeks before birth: NR
Entire pregnancy: NR
First 30 days of pregnancy: NR
Last 30 days of pregnancy: NR
T1:NR
T3:NR
T1: 0.97 (0.81, 1.1 8)°
T2: 1.06 (0.87, 1.27)d
T3:1.01 (0.83-1 .24)d
T1 : -0.9 g (-6.6, 4.8)
T2: -7.5 g (-13.8, 1.2)
T3: -4.5 g (-10.8, 1.8)
Last 30 days:
-1.1 g (-5.6, 3.4)
T1 : 2.8 g (-10.5, 16.0)
T2: 4.4 g (-11. 4, 20.1)
T3: 11. 3 g (-4.4, 27.1)
T1 : -3.2 g (-25.6, 19)
T2: -0.2 g (-23.8, 23.4)
T3: -6.0 g (-30.8, -18.8)
Entire pregnancy:
1.13(0.92,1.38)°
T1: 1.02 (0.85, 1.22)°
T2: 0.93 (0.78, 1.12)°
T3: 1 .05 (0.87, 1 .26)°
T1 : 0.87 (0.81 , 0.94)°
T3: 1.05 (1.02, 1.08)°
"Change in birthweight per 10 ppb change in 03
"Median
°0dds ratios of LEW; Highest quartile of exposure compared to lowest quartile of exposure
dRelative risk of LEW per 10 ppb change in 03
T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester
NR: No quantitative results reported
Table 7-5 provides a brief overview of the epidemiologic studies of birth weight. In
summary, only the Children's Health Study conducted in southern California (Salam et
al.. 2005) provides strong evidence for an effect of ambient O3 on birth weight. The study
by Morello-Frosch et al. (2010). also conducted in California, provides support for the
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1 results of the Children's Health Study. Additional studies conducted in the U.S., Canada,
2 Australia, Latin America, and Asia provide limited and inconsistent evidence to support
3 the effect reported in the Children's Health Study. The toxicological literature on the
4 effect of O3 on birth weight is sparse. In some studies, the reporting of birth weight may
5 be avoided because birth weight can be confounded by decreased litter size resulting
6 from an increased rate of pup resorption (aborted pups) in O3 exposed dams. In one
7 toxicological study by Haro and Paz (1993). no differences in litter size were observed
8 and decreased birth weight in pups from dams who were exposed to Ippm O3 during
9 pregnancy was reported. A second animal toxicology study recapitulated these finding
10 with pregnant BALB/c mice that exposed to O3 (1.2 ppm, GD9-18) producing pups with
11 significantly decreased birth weights (Sharkhuu et al.. 2011).
7.4.4 Preterm Birth
12 Preterm birth (PTB) is a syndrome (Romero et al., 2006) that is characterized by multiple
13 etiologies. It is therefore unusual to be able to identify an exact cause for each PTB. In
14 addition, PTB is not an adverse outcome in itself, but an important determinant of health
15 status (i.e., neonatal morbidity and mortality). Although some overlap exists for common
16 risk factors, different etiologic entities related to distinct risk factor profiles and leading
17 to different neonatal and postneonatal complications are attributed to PTB and measures
18 of fetal growth. Although both restricted fetal growth and PTB can result in LEW,
19 prematurity does not have to result in LEW or growth restricted babies.
20 A major issue in studying environmental exposures and PTB is selecting the relevant
21 exposure period, since the biological mechanisms leading to PTB and the critical periods
22 of vulnerability are poorly understood (Bobak. 2000). Exposures proximate to the birth
23 may be most relevant if exposure causes an acute effect. However, exposure occurring in
24 early gestation might affect placentation, with results observable later in pregnancy, or
25 cumulative exposure during pregnancy may be the most important determinant. The
26 studies reviewed have dealt with this issue in different ways. Many have considered
27 several exposure metrics based on different periods of exposure. Often the time periods
28 used are the first month (or first trimester) of pregnancy and the last month (or 6 weeks)
29 prior to delivery. Using a time interval prior to delivery introduces an additional problem
30 since cases and controls are not in the same stage of development when they are
31 compared. For example, a preterm infant delivered at 36 weeks is a 32-week fetus
32 4 weeks prior to birth, while an infant born at term (40 weeks) is a 36-week fetus 4 weeks
33 prior to birth.
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1 Recently, investigators have examined the association of PTB with both short-term (i.e.,
2 hours, days, or weeks) and long-term (i.e., months or years) exposure periods. Time-
3 series studies have been used to examine the association between air pollution
4 concentrations during the days immediately preceding birth. An advantage of these time-
5 series studies is that this approach can remove the influence of covariates that vary across
6 individuals over a short period of time. Retrospective cohort and case-control studies
7 have been used to examine long-term exposure periods, often averaging air pollution
8 concentrations over months or trimesters of pregnancy.
9 Reported studies fail to show consistency in pollutants and periods during pregnancy
10 when an effect occurs. For example, while some studies find the strongest effects
11 associated with exposures early in pregnancy, others report effects when the exposure is
12 limited to the second or third trimester. However, the effect of air pollutant exposure
13 during pregnancy on PTB has a biological basis. There is an expanding list of possible
14 mechanisms that may explain the association between O3 exposure and PTB (see
15 Section 5. 4.2.4).
16 Many studies of PTB compare exposure in quartiles, using the lowest quartile as the
17 reference (or control) group. No studies use a truly unexposed control group. If exposure
18 in the lowest quartile confers risk, than it may be difficult to demonstrate additional risk
19 associated with a higher quartile. Thus negative studies must be interpreted with caution.
20 Preterm birth occurs both naturally (idiopathic preterm), and as a result of medical
21 intervention (iatrogenicpreterm). Ritz et al. (2007; 2000) excluded all births by Cesarean
22 section to limit their studies to idiopathic preterm. No other studies attempted to
23 distinguish the type of PTB, although air pollution exposure maybe associated with only
24 one type. This is a source of potential effect misclassification.
25 Generally, studies of air pollution-birth outcome conducted in North America and the
26 United Kingdom have not identified an association between PTB and maternal exposure
27 to O3. Most recently, Darrow et al. (2009) used vital record data to construct a
28 retrospective cohort of 476,489 births occurring between 1994 and 2004 in 5 central
29 counties of metropolitan Atlanta. Using a time-series approach, the authors examined
30 aggregated daily counts of PTB in relation to ambient levels of CO, NO2, SO2, O3, PM10,
31 PM2 5 and speciated PM measurements. This study investigated 3 gestational windows of
32 exposure: the first month of gestation, the final week of gestation, and the final 6 weeks
33 of gestation. The authors did not observe associations of PTB with O3.
34 A number of U.S. studies were conducted in southern California, and report somewhat
35 inconsistent results. Ritz et al. (2000) evaluated the effect of air pollution (CO, NO2, O3,
36 PMio) exposure during pregnancy on the occurrence of PTB in a cohort of 97,518
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1 neonates born in southern California between 1989 and 1993. The authors use both short-
2 and long-term exposure windows, averaging pollutant measures taken at the closest air-
3 monitoring station over distinct periods, such as 1, 2, 4, 6, 8, 12, and 26 weeks before
4 birth and the whole pregnancy period. Additionally, they calculated average exposures
5 for the first and second months of pregnancy. The authors found no consistent effects for
6 O3 over any of the pregnancy periods in single or multi-pollutant models. Wilhelm and
7 Ritz (2005) extended previous analyses of PTB (Ritz et al., 2000; Ritz and Yu. 1999) in
8 California to include 1994-2000. The authors examined varying residential distances
9 from monitoring stations to see if the distance affected risk estimation, because effect
10 attenuation may result from local pollutant heterogeneity inadequately captured by
11 ambient monitors. The authors analyzed the association between O3 exposure during
12 varying periods of pregnancy and PTB, finding a positive association between O3 levels
13 in both the first trimester of pregnancy (RR=1.23 [95% CI: 1.06, 1.42] per 10 ppb
14 increase in 24-h avg O3) and the first month of pregnancy (results for first trimester
15 exposure were similar, but slightly smaller, quantitative results not presented) in models
16 containing all pollutants. No association was observed between O3 in the 6 weeks before
17 birth and preterm delivery. Finally, Ritz et al. (2007) conducted a case-control survey
18 nested within a birth cohort and assessed the extent to which residual confounding and
19 exposure misclassification impacted air pollution effect estimates. The authors calculated
20 mean exposure levels for three gestational periods: the entire pregnancy, the first
21 trimester, and the last 6 weeks before delivery. Though positive associations were
22 observed for CO and PM2 5, no consistent patterns of increase in the odds of PTB for O3
23 or NO2 were observed.
24 One study conducted in Canada evaluated the impacts of air pollution (including CO,
25 NO2, NO, O3, SO2, PM25, and PM10) on PTBs (1999-2002) using spatiotemporal
26 residential exposure metrics by month of pregnancy in Vancouver, BC (Brauer et al..
27 2008). The authors did not observe consistent associations with any of the pregnancy
28 average exposure metrics except for PM25 for PTB. The O3 associations were largely
29 protective, and no quantitative results were presented for O3. Additionally, Lee et al.
30 (2008c) used time-series techniques to investigate the short-term associations of O3 and
31 PTB in London, England. In addition to exposure on the day of birth, cumulative
32 exposure up to 1 week before birth was investigated. The risk of PTB did not increase
33 with exposure to the levels of ambient air pollution experienced by this population.
34 Conversely, studies conducted in Australia and China provide evidence for an association
35 between ambient O3 and PTB. Hansen et al. (2006) reported that exposure to O3 during
36 the first trimester was associated with an increased risk of PTB (OR=1.38, [95% CI:
37 1.14, 1.69] per 10 ppb increase). Although the test for trend was significant due to the
38 strong effect in the highest quartile, there was not an obvious exposure-response pattern
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1 across the quartiles of O3 during the first trimester. The effect estimate was diminished
2 and lost statistical significance when PMi0 was included in the model (OR=1.23, [95%
3 CI: 0.97, 1.59] per 10 ppb increase). Maternal exposure to O3 during the 90 days prior to
4 birth showed a weak, positive association with PTB (OR=1.09, [95% CI: 0.85, 1.39] per
5 10 ppb increase). Jalaludin et al. (2007) found that O3 levels in the month and
6 three months preceding birth had a statistically significant association with PTB. Ozone
7 levels in the first trimester of pregnancy were associated with increased risks for PTBs
8 (OR=1.15 [95% CI: 1.05, 1.24] per 10 ppb increase in 1-h max O3 concentration), and
9 remained a significant predictor of PTB in copollutant models (ORs between 1.07 and
10 1.10). ORs increased for first month of pregnancy when restricted to within 5 km of a
11 monitoring station (OR=1.60, [95% CI: 1.27, 2.03]), but did not show a cumulative effect
12 for first 3 months of pregnancy (OR=0.81, [95% CI: 0.67, 0.98]). Jiang et al. (2007)
13 examined the acute effect of air pollution on PTB, including risk in relation to levels of
14 pollutants for a single day exposure window with lags from 0 to 6 days before birth. An
15 increase of 10 ppb of the 8-week avg of O3 corresponded to 9.47 % (95% CI: 0.70,
16 18.7%) increase in PTBs. Increases in PTB were also observed for PMi0, SO2, and NO2.
17 The authors did not observe any significant acute effect of outdoor air pollution on PTB
18 among the 1-day acute time windows examined in the week before birth.
19 Little data is available from toxicological studies; one study reported a nearly statistically
20 significant increase in pregnancy duration in mice when exposed to 0.8 or 1.2 ppm O3.
21 This phenomenon was most likely due to the anorexigenic effect of relatively high O3
22 concentrations (Bignami et al.. 1994).
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1
2
3
4
Table 7-6
Study
Darrow et al. (2009)
Ritzetal. (2000)
Wilhelm and Ritz (2005)
Ritzetal. (2007)
Braueretal. (2008)
Lee et al. (2008c)
Hansen et al. (2006)
Jalaludin et al. (2007)
Jiang etal. (2007)
Brief summary of epidemiologic studies of PTB
Location
Sample Size
Atlanta, GA
(n=476,489)
California, US
(n=97,158)
Los Angeles, CA
(n=1 06,483)
Los Angeles, CA
(n=58,316)
Vancouver, BC,
Canada
(n=70,249)
London, UK
Brisbane, Australia
(n=28,200)
Sydney, Australia
(n=1 23,840)
Shanghai, China
(n=3,346 preterm
births)
Mean O3
(PPb)
8-h max: 44.1
8 h: 36.9
1 h: 21. 1-22.2
24-h avg: 22.5
24-h avg:
14
24-h avg: NR
8-h max:
26.7
1-h max:
30.9
8-h avg:
32.7
Exposure
assessment
Population-weighted
spatial averages
Nearest Monitor (within
4 miles)
<2 mi of monitor
Varying distances to
monitor
Nearest monitor to ZIP
code
Nearest Monitor (within
10km)
Inverse Distance
Weighting (IDW)
1 monitor
City-wide avg
City-wide avg and <5
km from monitor
City-wide avg
Effect Estimate3 (95% Cl)
First month: 0.98 (0.97, 1.00)
Last week: 0.99 (0.98, 1 .00)
Last 6 weeks: 1.00 (0.98, 1.02)
First month: NR
Last 6 weeks: NR
First month: 1.23 (1.06, 1.42)
T1:NR
12:1.38(1.14,1.66)
Last 6 weeks: NR
Entire pregnancy: NR
11:0.93(0.82,1.06)
Last 6 weeks: NR
Entire pregnancy: NR
First 30 days of pregnancy: NR
Last 30 days of pregnancy: NR
T1:NR
T3:NR
Lag 0:1. 00 (1.00, 1.01)
11:1.39(1.15,1.70)
T3: 1 .09 (0.88, 1 .39)
First month: 1.604 (1.268,2.030)"
T1 : 0.807 (0.668, 0.976)b
T3: 1.011 (0.910,1.124)"
Last month: 0.984 (0.906, 1 .069)"
4 wks before birth: 1 .06 (1 .00, 1 .12)
6 wks before birth: 1 .06 (0.99, 1 .13)
8 wks before birth: 1 .09 (1 .01 , 1 .19)
LO: NR (results presented in figure)
L1: NR (results presented in figure)
L2: NR (results presented in figure)
L3: NR (results presented in figure)
L4: NR (results presented in figure)
L5: NR (results presented in figure)
L6: NR (results presented in figure)
"Relative risk of PTB per 10 ppb change in 03.
"Relative risk of PTB per 1 ppb change in 03.
T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester
LO = Lag 0, L1= Lag 1, L2 = Lag 2, L3 = Lag 3, L4 = Lag 4, L5 = Lag 5, L6 = Lag 6
NR: No quantitative results reported
Table 7-6 provides a brief overview of the epidemiologic studies of PTB. In summary,
the evidence is consistent when examining shorter-term, late-pregnancy exposure to O3
and reports no association with PTB. However when long-term exposure to O3 early in
pregnancy is examined the results are inconsistent. Studies conducted in the U.S.,
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1 Canada, and England find no association with O3 and PTB, while studies conducted in
2 Australia and China report an O3 effect on PTB.
7.4.5 Fetal Growth
3 Low birth weight has often been used as an outcome measure because it is easily
4 available and accurately recorded on birth certificates. However, LEW may result from
5 either short gestation, or inadequate growth in utero. Most of the studies investigating air
6 pollution exposure and LEW limited their analyses to term infants to focus on inadequate
7 growth. A number of studies were identified that specifically addressed growth restriction
8 in utero by identifying infants who failed to meet specific growth standards. Usually
9 these infants had birth weight less than the 10th percentile for gestational age, using an
10 external standard. Many of these studies have been previously discussed, since they also
11 examined other reproductive outcomes (i.e., LEW or PTB).
12 Fetal growth is influenced by maternal, placental, and fetal factors. The biological
13 mechanisms by which air pollutants may influence the developing fetus remain largely
14 unknown. Several mechanisms have been proposed, and are the same as those
15 hypothesized for birth weight (see Section 5. 4.2.4). Additionally, in animal toxicology
16 studies, O3 causes transient anorexia in exposed pregnant dams. This may be one of
17 many possible contributors to O3-dependent decreased fetal growth.
18 A limitation of environmental studies that use birth weight as a proxy measure of fetal
19 growth is that patterns of fetal growth during pregnancy cannot be assessed. This is
20 particularly important when investigating pollutant exposures during early pregnancy as
21 birth weight is recorded many months after the exposure period. The insult of air
22 pollution may have a transient effect on fetal growth, where growth is hindered at one
23 point in time but catches up at a later point. For example, maternal smoking during
24 pregnancy can alter the growth rate of individual body segments of the fetus at variable
25 developmental stages, as the fetus experiences selective growth restriction and
26 augmentation (Lampl and Jeanty. 2003).
27 The terms small-for-gestational-age (SGA), which is defined as a birth weight <10th
28 percentile for gestational age (and often sex and/or race), and intrauterine growth
29 retardation (IUGR) are often used interchangeably. However, this definition of SGA does
30 have limitations. For example, using it for IUGR may overestimate the percentage of
31 "growth-restricted" neonates as it is unlikely that 10% of neonates have growth
32 restriction (Wollmann. 1998). On the other hand, when the 10th percentile is based on the
33 distribution of live births at a population level, the percentage of SGA among PTB is
34 most likely underestimated (Hutcheon and Platt 2008). Nevertheless, SGA represents a
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1 statistical description of a small neonate, whereas the term IUGR is reserved for those
2 with clinical evidence of abnormal growth. Thus all IUGR neonates will be SGA, but not
3 all SGA neonates with be IUGR (Wollmann. 1998). In the following section the terms
4 SGA and IUGR are referred to as each cited study used the terms.
5 Over the past decade a number of studies examined various metrics of fetal growth
6 restriction. Salam et al. (2005) assessed the effect of increasing O3 concentrations on
7 IUGR in a population of infants born in California from 1975-1987 as part of the
8 Children's Health Study. The authors reported that maternal O3 exposures averaged over
9 the entire pregnancy and during the third trimester were associated with increased risk of
10 IUGR. A 10-ppb difference in 24-h maternal O3 exposure during the third trimester
11 increased the risk of IUGR by 11% (95% CI: 0, 20%). Brauer et al. (2008) evaluated the
12 impacts of air pollution (CO, NO2, NO, O3, SO2, PM2 5, PM10) on SGA (1999-2002)
13 using spatiotemporal residential exposure metrics by month of pregnancy in Vancouver,
14 BC. The O3 associations were largely protective (OR= 0.87, [95% CI: 0.81, 0.93] for a
15 10 ppb increase in inverse distance weighted SGA), and no additional quantitative results
16 were presented for O3. Liu et al. (2007b) examined the association between IUGR among
17 singleton term live births and SO2, NO2, CO, O3, and PM2 5 in 3 Canadian cities for the
18 period 1985-2000. No increase in the risk of IUGR in relation to exposure to O3 averaged
19 over each month and trimester of pregnancy was noted.
20 Three studies conducted in Australia provide evidence for an association between
21 ambient O3 and fetal growth restriction. Hansen et al. (2007) examined SGA among
22 singleton, full-term births in Brisbane, Australia in relation to ambient air pollution (bsp,
23 PMio, NO2, O3) during pregnancy. They also examined head circumference and crown-
24 heel length in a subsample of term neonates. Trimester specific exposures to all pollutants
25 were not statistically significantly associated with a reduction in head circumference or
26 an increased risk of SGA. When monthly-specific exposures were examined, the authors
27 observed an increased risk of SGA associated with exposure to O3 during month 4
28 (OR=1.11 [95% CI: 1.00, 1.24] per 10 ppb increase). In a subsequent study, Hansen et al.
29 (2008) examined the possible associations between fetal ultrasonic measurements and
30 ambient air pollution (PM10, O3, NO2, SO2) during early pregnancy. This study had two
31 strengths: (1) fetal growth was assessed during pregnancy as opposed to at birth; and (2)
32 there was little delay between exposures and fetal growth measurements, which reduces
33 potential confounding and uses exposures that are concurrent with the observed growth
34 pattern of the fetus. Fetal ultrasound biometric measurements were recorded for biparietal
35 diameter (BPD), femur length, abdominal circumference, and head circumference. To
36 further improve exposure assessment, the authors restricted the samples to include only
37 scans from women for whom the centroid of their postcode was within 14 km of an air
38 pollution monitoring site. Ozone during days 31-60 was associated with decreases in all
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1 of the fetal growth measurements, and a 1.78 mm reduction in abdomen circumference
2 per 10 ppb increase in O3 concentration, though this effect did not persist in copollutant
3 models. The change in ultrasound measurements associated with O3 during days 31-60 of
4 gestation indicated that increasing O3 concentration decreased the magnitude of
5 ultrasound measurements for women living within 2 km of the monitoring site. The
6 relationship decreased toward the null as the distance from the monitoring sites increased.
7 When assessing effect modification due to SES, there was some evidence of effect
8 modification for most of the associations, with the effects of air pollution stronger in the
9 highest SES quartile. In the third study, Marines et al. (2005) estimated the effects of
10 pollutant (PMio, PM2 5, NO2, CO and O3) exposure in the first, second and third
11 trimesters of pregnancy and risk of SGA in Sydney, Australia. Citywide average air
12 pollutant concentrations in the last month, third trimester, and first trimester of pregnancy
13 had no effect on SGA. Concentrations of O3 in the second trimester of pregnancy had
14 small but adverse effects on SGA (OR=1.10 [95% CI: 1.00, 1.14] per 10 ppb increment).
15 This effect disappeared when the analysis was limited to births with a maternal address
16 within 5 km of a monitoring station (OR=1.00 [95% CI: 0.60, 1.79] per 10 ppb
17 increment).
18 Very little information from toxicological studies is available to address effects on fetal
19 growth. However, there is evidence to suggest that prenatal exposure to O3 can affect
20 postnatal growth. A few studies reported that mice or rats exposed developmentally
21 (gestationally ± lactationally) to O3 had deficits in body weight gain in the postpartum
22 period (Bignami et al.. 1994: Haro and Paz. 1993: Kavlock et al.. 1980).
23 Table 7-7 provides a brief overview of the epidemiologic studies of fetal growth
24 restriction. In summary, the evidence is inconsistent when examining exposure to O3 and
25 fetal growth restriction. Similar to PTB, studies conducted in Australia have reported an
26 effect of O3 on fetal growth, whereas studies conducted in other areas have not found
27 such an effect. This may be due to the restriction of births to those within 2-14 km of a
28 monitoring station, as was done in the Australian studies.
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Table 7-7
Study
Salametal.
(2005)
Braueretal.
(2008)
Liu et al. (2007b)
Hansen et al.
(2007)
Hansen et al.
(2008)
Mannes etal.
(2005)
Brief summary of epidemiologic studies of fetal growth
Location
(Sample Size)
California, U.S.
(n=3901)
Vancouver, BC, Canada
(n=70,249)
Calgary, Edmonton, and
Montreal, Canada
(n= 16,430)
Brisbane, Australia
(n=26,617)
Brisbane, Australia
(n=15,623)
Sydney, Australia
(n=138,056)
Mean O3 (ppb)
24-h avg:
27.3
8h:
50.6
24-h avg:
14
24-h avg:
16.5
1-h max:
31.2
8-h max:
26.7
8-h avg:
24.8
1-h max:
31.6
Exposure
assessment
ZIP code level
Nearest Monitor (within
10km)
Inverse Distance
Weighting (IDW)
Census Subdivision avg
City-wide avg
Within 2 km of monitor
City-wide avg and
<5 km from monitor
Effect Estimate3 (95% Cl)
Entire pregnancy: 1 .16 (1 .00, 1 .32)
11:1.00(0.94,1.11)
12:1.06(1.00,1.12)
73:1.11 (1.00,1.17)
Entire pregnancy: NR
First 30 days of pregnancy: NR
Last 30 days of pregnancy: NR
71:NR
73: NR
Entire pregnancy: NR (results presented in
figure)
71 : NR (results presented in figure)
72: NR (results presented in figure)
73: NR (results presented in figure)
71:1.01 (0.89,1.15)
72:1.00(0.86,1.17)
73: 0.83 (0.71 , 0.97)
M1: -0.32 (-1.56,0.91)"
M2: -0.58 (-1.97,0.80)"
M3: 0.26 (-1.07, 1.59)"
M4:0.11 (-0.98,1.21)"
71 : 0.90 (0.48, 1 .34)
72:1.00(0.60,1.79)
73:1.10(0.66,1.97)
Last 30 days of pregnancy: 1 .10 (0.74,
1.79)
'Relative risk of fetal growth restriction per 10 ppb change in 03.
"Mean change in fetal ultrasonic measure of head circumference recorded between 13 and 26 weeks gestation for a 10-ppb increase in maternal
exposure to 03 during early pregnancy
71 = First 7rimester, 72 = Second 7rimester, 73 = 7hird 7rimester
M1 = Month 1, M2 = Month 2, M3 = Month 3, M4 = Month 4
NR: No quantitative results reported
1
2
3
4
5
6
7.4.6 Postnatal growth
Time-pregnant BALB/c mice were exposed to O3 (0, 0.4, 0.8, or 1.2 ppm) GD9-18 with
parturition at GD20-21 (Sharkhuu et al.. 2011). As the offspring aged, postnatal litter
body weight continued to be significantly decreased in the highest dose (1.2 ppm) O3
group at PND3 and PND7. When the pups were weighed separately by sex at PND42, the
males with the highest dose of O3 exposure (1.2 ppm, GD9-18) had significant
decrements in body weight (Sharkhuu et al.. 2011).
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7.4.7 Birth Defects
1 Despite the growing body of literature evaluating the association between ambient air
2 pollution and various adverse birth outcomes, relatively few studies have investigated the
3 effect of temporal variations in ambient air pollution on birth defects. Heart defects and
4 oral clefts have been the focus of the majority of these recent studies, given the higher
5 prevalence than other birth defects and associated mortality. Mechanistically, air
6 pollutants could be involved in the etiology of birth defects via a number of key events
7 (see Section 5. 4.2.4).
8 Several studies have been conducted examining the relationship between O3 exposure
9 during pregnancy and birth defects and reported a positive association with cardiac
10 defects. The earliest of these studies was conducted in southern California (Ritz et al..
11 2002). This study evaluated the effect of air pollution on the occurrence of cardiac birth
12 defects in neonates and fetuses delivered in southern California in 1987-1993. Maternal
13 exposure estimates were based on data from the fixed site closest to the mother's ZIP
14 code area. When using a case-control design where cases were matched to 10 randomly
15 selected controls, results showed increased risks for aortic artery and valve defects
16 (OR=1.56 [95% CI: 1.16, 2.09] per 10 ppb O3), pulmonary artery and valve anomalies
17 (OR=1.34 [95% CI: 0.96, 1.87] per 10 ppb O3), and conotruncal defects (OR=1.36 [95%
18 CI: 0.91, 2.03] per 10 ppb O3) in a dose-response manner with second-month O3
19 exposure. A study conducted in Texas (Gilboaetal.. 2005) looked at a similar period of
20 exposure but reported no association with most of the birth defects studied (O3
21 concentration was studied using quartiles with the lowest representing <18 ppb and the
22 highest representing > 31 ppb). The authors found slightly elevated odds ratios for
23 pulmonary artery and valve defects. They also detected an inverse association between
24 O3 exposure and isolated ventricular septal defects. Overall, this study provided some
25 weak evidence that air pollution increases the risk of cardiac defects. Hansen et al. (2009)
26 investigated the possible association between ambient air pollution and the risk of cardiac
27 defects. When analyzing all births with exposure estimates for O3 from the nearest
28 monitor there was no indication for an association with cardiac defects. There was also
29 no adverse association when restricting the analyses to only include births where the
30 mother resided within 12 km of a monitoring station. However, among births within 6 km
31 of a monitor, a 10 ppb increase in O3 was associated with an increased risk of pulmonary
32 artery and valve defects (OR=8.76 [95% CI: 1.80, 56.55]). As indicated by the very wide
33 credible intervals, there were very few cases in the sensitivity analyses for births within 6
34 km of a monitor, and this effect could be a result of type I errors. Dadvand et al. (2011)
3 5 investigated the association between maternal exposure to ambient air pollution and the
36 occurrence of cardiac birth defects in England. Similar to Hansen et al. (2009), they
37 found no associations with maternal exposure to O3 except for when the analysis was
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1 limited to those subjects residing within a 16 km distance of a monitoring station (OR for
2 malformations of pulmonary and tricuspid valves=1.64 [95% CI: 1.04, 2.60] per 10 ppb
3 increase in O3).
4 Despite the association between O3 and cardiac defects observed in the above studies, a
5 recent study did not observe an increased risk of cardiac birth defects associated with
6 ambient O3 concentrations. The study, conducted in Atlanta, GA, examined O3 exposure
7 during the third through seventh week of pregnancy and reported no association with risk
8 of cardiovascular malformations (mean long-term average of 8-h O3 concentrations
9 excluding November through February ranged by 5-year groups from 39.8 to 43.3 ppb)
10 (Strickland etaL 2009).
11 Several of these studies have also examined the relationship between O3 exposure during
12 pregnancy and oral cleft defects. The study by Ritz et al. (Ritz et al.. 2002) evaluated the
13 effect of air pollution on the occurrence of orofacial birth defects and did not observe
14 strong associations between ambient O3 concentration and orofacial defects. They did
15 report an OR of 1.13 (95% CI: 0.90, 1.40) per 10 ppb during the second trimester for cleft
16 lip with or without cleft palate. Similarly, Gilboa et al. (Gilboa et al.. 2005) reported and
17 OR of 1.09 (95% CI: 0.70, 1.69) for oral cleft defects when the fourth quartile was
18 contrasted with the first quartile of exposure during 3-8 weeks of pregnancy. Hansen et
19 al. (2009) reported no indication for an association with cleft defects. Hwang and Jaakola
20 (2008) conducted a population-based case-control study to investigate exposure to
21 ambient air pollution and the risk of cleft lip with or without cleft palate in Taiwan. The
22 risk of cleft lip with or without cleft palate was increased in relation to O3 levels in the
23 first gestational month (OR= 1.17 [95% CI: 1.01, 1.36] per 10 ppb) and second gestational
24 month (OR=1.22 [95% CI: 1.03, 1.46] per 10 ppb), but was not related to any of the other
25 pollutants. In three-pollutant models, the effect estimates for O3 exposure were stable for
26 the four different combinations of pollutants and were all statistically significant.
27 Marshall et al. (2010) compared estimated exposure to ambient pollutants during early
28 pregnancy among mothers of children with oral cleft defects to that among mothers of
29 controls. The authors observed no consistent elevated associations between any of the air
30 pollutants examined and cleft malformations, though there was a weak association
31 between cases of cleft palate only and increasing O3 concentrations. This association
32 increased when cases and controls were limited to those with residences within 10 km of
33 the closest O3 monitor (OR=2.2 [95% CI: 1.0, 4.9], comparing highest quartile [>33 ppb]
34 to lowest quartile [<15 ppb]).
35 A limited number of toxicological studies have examined birth defects in animals
36 exposed gestationally to O3. Kavlock et al. (1979) exposed pregnant rats to O3 for precise
37 periods during organogenesis. No significant teratogenic effects were found in rats
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
exposed 8 hr/day to concentrations of O3 varying from 0.44 to 1.97 ppm during early
(days 6-9), mid (days 9-12), or late (days 17 to 20) gestation, or the entire period of
organogenesis (days 6-15). Earlier research found eyelid malformation following
gestational and postnatal exposure to 0.2 ppm O3 (Veninga. 1967).
Table 7-8 provides a brief overview of the epidemiologic studies of birth defects. These
studies have focused on cardiac and oral cleft defects, and the results from these studies
are not entirely consistent. This inconsistency could be due to the absence of true
associations between O3 and risks of cardiovascular malformations and oral cleft defects;
it could also be due to differences in populations, pollution levels, outcome definitions, or
analytical approaches. The lack of consistency of associations between O3 and
cardiovascular malformations or oral cleft defects might be due to issues relating to
statistical power or measurement error. A recent meta-analysis of air pollution and
congenital anomalies concluded that there was no statistically significant increase in risk
of congenital anomalies and O3 (Vrijheid et al., 2011). These authors note that
heterogeneity in the results of these studies may be due to inherent differences in study
location, study design, and/or analytic methods, and comment that these studies have not
employed some recent advances in exposure assessment used in other areas of air
pollution research that may help refine or reduce this heterogeneity.
Table 7-8
Study
Ritzetal. (2002)
Gilboaetal. (2005)
Hwang and Jaakola
Strickland etal. (2009)
Hansen et al. (2009)
Marshall et al. (2010)
Dadvand et al. (2011)
Brief summary of epidemiologic studies of birth defects
ExlCmineedS
Cardiac and Cleft
Defects
Cardiac and Cleft
Defects
Oral Cleft Defects
Cardiac Defects
Cardiac and Cleft
Defects
Oral Cleft Defects
Cardiac Defects
(Samp^Size)
Southern California
(n=3,549 cases;
10,649 controls)
7 Counties in TX
(n=5,338 cases;
4,580 controls)
Taiwan
(n=653 cases;
6,530 controls)
Atlanta, GA
(n=3,338 cases)
Brisbane, Australia
(n=150,308 births)
New Jersey
(n=71 7 cases;
12,925 controls)
Northeast England
(n=2, 140 cases;
14,256 controls)
Mean 03 (p
24-h avg:
NR
24-h avg:
NR
24-h avg:
27.31
8-h max:
39.8-43.3
8-h max:
25.8
24-h avg:
25
24-h avg:
18.8
.. Exposure
p ' Assessment
Nearest Monitor
(within 10 mi)
Nearest Monitor
Inverse Distance
Weighting (IDW)
Weighted City-wide
avg
Nearest Monitor
Nearest Monitor
(within 40 km)
Nearest Monitor
Exposure Window
Month 1,2,3
Trimester 2,3
3-mo period prior to
conception
Weeks 3-8 of gestation
Months 1,2,3
Weeks 3-7 of gestation
Weeks 3-8 of gestation
Weeks 5-1 0 of gestation
Weeks 3-8 of
gestation' 1
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7.4.8 Developmental Respiratory Effects
1 The issue of prenatal exposure has assumed increasing importance since ambient air
2 pollution exposures of pregnant women have been shown to lead to adverse pregnancy
3 outcomes, as well as to respiratory morbidity and mortality in the first year of life.
4 Growth and development of the respiratory system take place mainly during the prenatal
5 and early postnatal periods. This early developmental phase is thought to be very
6 important in determining long-term lung growth. Studies have recently examined this
7 emerging issue. Several studies were included in Sections 7.2.1 and 7.2.3, and are
8 included here because they reported both prenatal and post-natal exposure periods.
9 Mortimer et al. (2008a, b) examined the association of prenatal and lifetime exposures to
10 air pollutants with pulmonary function and allergen sensitization in a subset of asthmatic
11 children (ages 6-11) included in the Fresno Asthmatic Children's Environment Study
12 (FACES). Monthly means of pollutant levels for the years 1989-2000 were created and
13 averaged separately across several important developmental time-periods, including the
14 entire pregnancy, each trimester, the first 3 years of life, the first 6 years of life, and the
15 entire lifetime. The 8-h avg O3 concentrations were approximately 50 ppb for each of the
16 exposure metrics (estimated from figure). In the first analysis (Mortimer et al.. 2008a).
17 negative effects on pulmonary function were found for exposure to PM10, NO2, and CO
18 during key neonatal and early life developmental periods. The authors did not find a
19 negative effect of exposure to O3 among this cohort. In the second analysis (Mortimer et
20 al.. 2008b). sensitization to at least one allergen was associated, in general, with higher
21 levels of CO and PM10 during the entire pregnancy and second trimester and higher PM10
22 during the first 2 years of life. Lower exposure to O3 during the entire pregnancy or
23 second trimester was associated with an increased risk of allergen sensitization. Although
24 the pollutant metrics across time periods are correlated, the strongest associations with
25 the outcomes were observed for prenatal exposures. Though it may be difficult to
26 disentangle the effect of prenatal and postnatal exposures, the models from this group of
27 studies suggest that each time period of exposure may contribute independently to
28 different dimensions of school-aged children's pulmonary function. For 4 of the 8
29 pulmonary-function measures (FVC, FEVi, PEF, FEF25-75), prenatal exposures were
30 more influential on pulmonary function than early-lifetime metrics, while, in contrast, the
31 ratio of measures (FEVi/FVC and FEF25_75/FVC) were most influenced by postnatal
32 exposures. When lifetime metrics were considered alone, or in combination with the
33 prenatal metrics, the lifetime measures were not associated with any of the outcomes,
34 suggesting the timing of the exposure may be more important than the overall dose and
35 prenatal exposures are not just markers for lifetime or current exposures.
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1 Clark et al. (2010) investigated the effect of exposure to ambient air pollution in utero
2 and during the first year of life on risk of subsequent asthma diagnosis (incident asthma
3 diagnosis up to age 3-4) in a population-based nested case-control study. Air pollution
4 exposure for each subject based on their residential address history was estimated using
5 regulatory monitoring data, land use regression modeling, and proximity to stationary
6 pollution sources. An average exposure was calculated for the duration of pregnancy
7 (-15 ppb) and the first year of life (-14 ppb). In contrast to the Mortimer et al. studies
8 (2008a. b), the effect estimates for first-year exposure were generally larger than for in
9 utero exposures. However, similar to the Mortimer et al. studies, the observed
10 associations with O3 were largely protective. Because of the relatively high correlation
11 between in utero and first-year exposures for many pollutants, it was difficult to discern
12 the relative importance of the individual exposure periods.
13 Latzin et al. (2009) examined whether prenatal exposure to air pollution was associated
14 with lung function changes in the newborn. Tidal breathing, lung volume, ventilation
15 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age=
16 5 weeks). The median of the 24-h avg O3 concentrations averaged across the post-natal
17 period was -44 ppb. Consistent with the previous studies, no association was found for
18 prenatal exposure to O3 and lung function.
19 The new toxicological literature since the 2006 O3 AQCD, covering respiratory changes
20 related to developmental O3 exposure, reports ultrastructural changes in bronchiole
21 development, alterations in placental and pup cytokines, and increased pup airway hyper-
22 reactivity. These studies are detailed below. Older studies are discussed where new
23 information is not available.
24 Fetal rat lung bronchiole development is triphasic, comprised of the glandular phase
25 (measured at GDIS), the canalicular phase (GD20), and the saccular phase (GD21). The
26 ultrastructural lung development in fetuses of pregnant rats exposed to 1-ppm O3 (12
27 h/day, out to either GDIS, GD20 or GD21) was examined by electron microscopy during
28 these three phases. In the glandular phase, bronchiolar columnar epithelial cells in fetuses
29 of dams exposed to O3 had cytoplasmic damage and swollen mitochondria. Bronchial
30 epithelium at the canalicular phase in O3 exposed pups had delayed maturation in
31 differentiation, i.e., glycogen abundance in secretory cells had not diminished as it should
32 with this phase of development. Congruent with this finding, delayed maturation of
33 tracheal epithelium following early neonatal O3 exposure (1 ppm, 4-5 h/day for
34 first week of life) in lambs has been previously reported (Mariassy et al.. 1990; Mariassy
35 et al., 1989). Also at the canalicular phase, atypical cells were seen in the bronchiolar
36 lumen of O3 exposed rat fetuses. Finally, in the saccular phase, mitochondrial
37 degradation was present in the non-ciliated bronchiolar cells of rats exposed in utero to
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1 O3. In conclusion, O3 exposure of pregnant rats produced ultra-structural damage to near-
2 term fetal bronchiolar epithelium (Lopez et al.. 2008).
3 Exposure of laboratory animals to multiple airborne pollutants can differentially affect
4 pup physiology. One study showed that exposure of C57BL/6 mouse dams to 0.48 mg
5 PM intratracheally twice weekly for 3 weeks during pregnancy augmented O3-induced
6 airway hyper-reactivity in juvenile offspring. Maternal PM exposure also significantly
7 increased placental cytokines above vehicle-instilled controls. Pup postnatal O3 exposure
8 (1 ppm 3 h/day, every other day, thrice weekly for 4 weeks) induced significantly
9 increased cytokine levels (IL-1(3, TNF-a, KC, and IL-6) in whole lung versus postnatal
10 air exposed groups; this was further exacerbated with gestational PM exposure (Auten et
11 al.. 2009).
12 A series of experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3
13 starting at one-month of age have examined the effect of O3 alone or in combination with
14 an inhaled allergen on morphology and lung function (Plopper et al.. 2007). Exposure to
15 O3 alone or allergen alone produced small but not statistically significant changes in
16 baseline airway resistance and airway responsiveness, but the combined exposure to both
17 O3 + antigen produced statistically significant and greater than additive changes in both
18 functional measurements. Additionally, cellular changes and significant structural
19 changes in the respiratory tract have been observed in infant rhesus monkeys exposed to
20 O3 (Fanucchi et al., 2006). A more detailed description of these studies can be found in
21 Section 7.2.3 (Pulmonary Structure and Function), with mechanistic information found in
22 Section 5.4.2.4.
23 Lung immunological response in O3 exposed pups was followed by analyzing BAL and
24 lung tissue. Sprague Dawley (SD) pups were exposed to a single 3h exposure of air or O3
25 (0.6 ppm) on PND 13 (Han etal.. 2011). Bronchoalveolar lavage (BAL) was performed
26 10 hours after the end of O3 exposure. BALF polymorphonuclear leukocytes (PMNs) and
27 total BALF protein were significantly elevated in O3 exposed pups. Lung tissue from O3
28 exposed pups had significant elevations of manganese superoxide dismutase (SOD)
29 protein and significant decrements of extra-cellular SOD protein.
30 Various immunological outcomes were followed in offspring after their pregnant dams
31 (BALB/c mice) were exposed gestationally to O3 (0, 0.4, 0.8, or 1.2 ppm, GD9-18)
32 (Sharkhuu et al., 2011). Delayed type hypersensitivity (DTH) was initiated with initial
33 BSA injection at 6 weeks of age and then challenge 7 days later. The normal edematous
34 response of the exposed footpad (thickness after BSA injection) was recorded as an
35 indicator of DTH. In female offspring, normal footpad swelling with BSA injection that
36 was seen in air exposed animals was significantly attenuated with O3 exposure (0.8 and
37 1.2 ppm O3), implying immune suppression of O3 exposure specifically in DTH.
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1 Humoral immunity was measured with the sheep red blood cell (SRBC) response.
2 Animals received primary immunization with SRBC and then blood was drawn for
3 SRBC IgM measurement. A SRBC booster was given 2 weeks later with blood collected
4 5 days after booster for IgG measurement. Maternal O3 exposure had no effect on
5 humoral immunity in the offspring as measured by IgG and IgM titers after SRBC
6 primary and booster immunizations (Sharkhuu et al.. 2011).
7 Toxicity assessment and allergen sensitization was also assessed in these O3 exposed
8 offspring. At PND42, animals were euthanized for analysis of immune and inflammatory
9 markers (immune proteins, inflammatory cells, T cell populations in the spleen). A subset
10 of the animals was intra-nasally instilled or sensitized with ovalbumin on either PND2
11 and 3 or PND42 and 43. All animals were challenged with OVA on PND54, 55, and 56.
12 One day after final OVA challenge, lung function, lung inflammation and immune
13 response were determined. Offspring of O3 exposed dams that were initially sensitized at
14 PDN3 (early) or PND42 (late) were tested to determine the level of allergic sensitization
15 or asthma-like inflammation after OVA challenge. Female offspring sensitized early in
16 life developed significant eosinophilia (1.2 ppm O3) and elevated serum OVA-specific
17 IgE (1.2 ppm O3), which is a marker of airway allergic inflammation. The females that
18 were sensitized early also had significant decrements in BALF total cells, macrophages,
19 and lymphocytes (1.2 ppm O3). Offspring that were sensitized later (PND42) in life did
20 not develop the aforementioned changes in BALF, but these animals did develop modest,
21 albeit significant neutropenia (0.8 and 1.2 ppm O3) (Sharkhuu et al.. 2011).
22 BALF cytology in non-sensitized animals was followed. BALF of offspring born to dams
23 exposed to O3 was relatively unaffected (cytokines, inflammatory cell numbers/types) as
24 were splenic T cell subpopulations. LDH was significantly elevated in BALF of females
25 whose mothers were exposed to 1.2 ppm during pregnancy (Sharkhuu et al.. 2011). In
26 summary, the females born to mothers exposed to O3 developed modest
27 immunocompromise. Males were unaffected (Sharkhuu et al.. 2011).
28 Overall, animal toxicological studies have reported ultrastructural changes in bronchiole
29 development, alterations in placental and pup cytokines, and increased pup airway hyper-
30 reactivity related to exposure to O3 during the developmental period. Epidemiologic
31 studies have found no association between prenatal exposure to O3 and growth and
32 development of the respiratory system. Fetal origins of disease have received a lot of
33 attention recently, thus additional research to further explore the inconsistencies between
34 these two lines of evidence is warranted.
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7.4.9 Developmental Central Nervous System Effects
1 The following sections describe the results of toxicological studies of O3 and
2 developmental central nervous system effects. No epidemiologic studies of this
3 association have been published.
7.4.9.1 Laterality
4 Two reports of laterality changes in mice developmentally exposed to O3 have been
5 reported in the literature. Mice developmentally exposed to 0.6 ppm O3 (6 days before
6 breeding to weaning at PND21) showed a turning preference (left turns) distinct from air
7 exposed controls (clockwise turns) (Dell'Omo et al.. 1995); in previous studies this
8 behavior in mice has been found to correlate with specific structural asymmetries of the
9 hippocampal mossy fiber projections (Schopke etal.. 1991). The 2006 O3 AQCD
10 evidence for the effect of O3 on laterality or handedness demonstrated that rats exposed
11 to O3 during fetal and neonatal life showed limited, gender-specific changes in
12 handedness after exposure to the intermediate dose of O3 (only seen in female mice
13 exposed to 0.6 ppm O3, and not in males at 0.6 ppm or in either sex of 0.3 or 0.9 ppm O3
14 with exposure from 6 days before breeding to PND26) (Petruzzi et al.. 1999).
7.4.9.2 Brain Morphology and Neurochemical Changes
15 The nucleus tractus solitarius (NTS), a medullary area of respiratory control, of adult
16 animals exposed prenatally to 0.5 ppm O3 (12h/day, ED5-ED20) had significantly less
17 tyrosine hydroxylase staining versus control (Boussouar et al.. 2009). Tyrosine
18 hydroxylase is the rate-limiting enzyme for dopamine synthesis and serves as a precursor
19 for catecholamine synthesis; thus, decreased staining is used as a marker of dopaminergic
20 or catecholaminergic cell or activity loss in these regions and thus functions in neuronal
21 plasticity. After physical restraint stress, control animals respond at the histological level
22 with Fos activation, a marker of neuronal activity, and tyrosine hydroxylase activation in
23 the NTS, a response which is absent or attenuated in adult animals exposed prenatally to
24 0.5 ppm O3 (Boussouar et al.. 2009) when compared to control air exposed animals who
25 also were restrained. The O3-exposed offspring in this study were cross-fostered to
26 control air exposed dams to avoid O3-dependent dam related neonatal effects on
27 offspring outcomes (i.e., dam behavioral or lactational contributions to pup outcomes)
28 (Boussouar etal.. 2009).
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1 Developmental exposure to 0.3 or 0.6 ppm O3 prior to mating pair formation through
2 GDI? induced significant increased levels of BDNF in the striatum of adult (PND140)
3 O3 exposed offspring as compared to control air exposed animals; these O3-exposed
4 animals also had significantly decreased level of NGF in the hippocampus versus control
5 (Santucci etal., 2006).
6 Changes in the pup cerebellum with prenatal 1 ppm O3 exposure include altered
7 morphology (Romero-Velazquez et al.. 2002; Rivas-Manzano and Paz. 1999). decreased
8 total area (Romero-Velazquez et al.. 2002). decreased number of Purkinje cells (Romero-
9 Velazquez et al.. 2002). and altered monoamine neurotransmitter content with the
10 catecholamine system affected and the indoleamine system unaffected by O3 (Gonzalez-
11 Pina et al.. 2008).
7.4.9.3 Neurobehavioral Outcomes
12 O3 administration to dams during pregnancy with or without early neonatal exposure has
13 been shown to contribute to multiple neurobehavioral outcomes in offspring that are
14 described in further detail below.
15 O3 administration (0.4, 0.8 or 1.2 ppm O3) during the majority of pregnancy (PD7-17) of
16 CD-I mice did not affect pup behavioral outcomes including early behavioral ultrasonic
17 vocalizations and more permanent later measurements (PND60 or 61) including pup
18 activity, habituation and exploration and d-amphetamine-induced hyperactivity (Bignami
19 et al.. 1994); these pups were all cross-fostered or reared on non- O3 exposed dams.
20 Testing for aggressive behavior in mice continuously exposed to O3 (0.3 or 0.6 ppm from
21 30 days prior to mating to GDI7) revealed that mice had significantly increased
22 defensive/ submissive behavior (increased freezing posturing on the first day only of a
23 multiple-day exam) versus air exposed controls (Santucci et al.. 2006). Similar to this and
24 as reported in previous AQCDs, continuous exposure of adult animals to O3 induced
25 significant increases in fear behavior and decreased aggression as measured by
26 significantly decreased freezing behavior (Petruzzi et al.. 1995).
27 Developmentally exposed animals also had significantly decreased amount of time spent
28 nose sniffing other mice (Santucci et al.. 2006): this social behavior deficit, decreased
29 sniffing time, was not found in an earlier study with similar exposures (Petruzzi et al..
30 1995). but sniffing of specific body areas was measured in Santucci et al. (2006) and total
31 number of sniffs of the entire body was measured in Petruzzi et al. (1995). The two
32 toxicology studies exploring social behavior (sniffing) employ different study designs
33 and find opposite effects in animals exposed to O3
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7.4.9.4 Sleep Aberrations after Developmental Ozone Exposure
1 The effect of gestational O3 exposure (1 ppm O3, 12h/day, during dark period) on sleep
2 patterns in rat offspring was followed using 24 h polysomnographic recordings at 30, 60
3 and 90 days of age (Haro and Paz. 1993). Ozone-exposed pups manifested with inverted
4 sleep-wake patterns or circadian rhythm phase-shift. Rat vigilance was characterized in
5 wakefulness, slow wave sleep (SWS), and paradoxical sleep (PS) using previously
6 characterized criteria. The O3 exposed offspring spent longer time in the wakefulness
7 state during the light period, more time in SWS during the period of darkness, and
8 showed significant decrements in PS. Chronic O3 inhalation significantly decreased the
9 duration of PS during both the light and dark periods (Haro and Paz. 1993). These effects
10 were consistent at all time periods measured (30, 60 and 90 days of age). These sleep
11 effects reported after developmental exposures expand upon the existing literature on
12 sleep aberrations in adult animals exposed to O3 [rodents: (Paz and Huitron-Resendiz.
13 1996; Aritoetal.. 1992); and cats: (Paz and Bazan-Perkins. 1992)]. A role for inhibition
14 of cyclooxygenase-2 and the interleukins and prostaglandins in the O3-dependent sleep
15 changes potentially exists with evidence from a publication on indomethacin pretreatment
16 attenuating O3-induced sleep aberrations in adult male animals (Rubio and Paz. 2003).
7.4.10 Early Life Mortality
17 Infants may be particularly susceptible to the adverse effects of air pollution. Within the
18 first year of life, infants develop rapidly; therefore their susceptibility may change within
19 weeks or months. During the neonatal and post-neonatal periods, the developing lung is
20 highly susceptible to environmental toxicants. The lung is not well developed at birth,
21 with 80% of alveoli being formed postnatally. An important question regarding the
22 association between O3 and infant mortality is the critical window of exposure during
23 development for which infants are susceptible. Several age intervals have been explored:
24 neonatal (<1 month); postneonatal (1 month to 1 year); and an overall interval for infants
25 that includes both the neonatal and postneonatal periods (<1 year). Within these various
26 age categories, multiple causes of deaths have been investigated, particularly total deaths
27 and respiratory-related deaths. The studies reflect a variety of study designs, exposure
28 periods, regions, and adjustment for confounders. As discussed below, a handful of
29 studies have examined the effect of ambient air pollution on neonatal and postneonatal
30 mortality, with the former the least studied. These studies varied somewhat with regard to
31 the outcomes and exposure periods examined and study designs employed.
32 A major issue in studying environmental exposures and infant mortality is selecting the
33 relevant exposure period, since the biological mechanisms leading to death and the
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1 critical periods of exposure are poorly understood. Both short-term (days to weeks) and
2 long-term (months to years) exposure studies are included in this section and are
3 characterized accordingly in the text and tables. All studies of infant mortality are
4 included in the Reproductive and Developmental Effects section, as opposed to the
5 sections devoted to all- and cause-specific mortality, because infant development
6 processes, much like fetal development processes, may be particularly susceptible to O3-
7 induced health effects. Exposures proximate to the death may be most relevant if
8 exposure causes an acute effect. However, exposure occurring in early life might affect
9 critical growth and development, with results observable later in the first year of life, or
10 cumulative exposure during the first year of life may be the most important determinant.
11 The studies reviewed below have dealt with this issue in different ways. Many have
12 considered several exposure metrics based on different periods of exposure.
7.4.10.1 Stillbirth
13 Pereira et al. (1998) investigated the association among daily counts of intrauterine
14 mortality (over 28 weeks of gestation) and air pollutant concentrations in Sao Paulo,
15 Brazil from 1991 through 1992. The association was strong for NO2, but lesser for SO2
16 and CO. These associations exhibited a short lag time, less than 5 days. No significant
17 association was detected between short-term O3 exposure and intrauterine mortality.
7.4.10.2 Infant Mortality, Less than 1 Year
18 Ritz et al. (2006) linked birth and death certificates for infants who died between 1989
19 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South
20 Coast Air Basin of California. The authors examined short- and long-term exposure
21 periods 2 weeks, 1 month, 2 months, and 6 months before each case subject's death and
22 reported no association between ambient levels of O3 and infant mortality. Similarly,
23 Diaz et al. (2004) analyzed the effects of extreme temperatures and short-term exposure
24 to air pollutants on daily mortality in children less than 1 year of age in Madrid, Spain,
25 from 1986 to 1997 and observed no statistically significant association between mortality
26 and O3 concentrations. Hajat et al. (2007) analyzed time-series data of daily infant
27 mortality counts in 10 major cities in the UK to quantify any associations with short-term
28 changes in air pollution. When the results from the 10 cities were combined there was no
29 relationship between O3 and infant mortality, even after restricting the analysis to just the
30 summer months.
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1 Conversely, a time-series study of infant mortality conducted in the southwestern part of
2 Mexico City in the years 1993-1995 found that infant mortality was associated with
3 short-term exposure to NO2 and O3 3-5 days before death, but not as consistently as with
4 PM. A 10-ppb increase in 24-h avg O3 was associated with a 2.78% increase (95% CI:
5 0.29, 5.26%) in infant mortality (lag 3) (Loomis et al., 1999). This increase was
6 attenuated, although still positive when evaluated in a two-pollutant model with PM2 5.
7 One-hour max concentrations of O3 exceeded prevailing Mexican and international
8 standards nearly every day.
7.4.10.3 Neonatal Mortality, Less than 1 Month
9 Several studies have evaluated ambient O3 concentrations and neonatal mortality and
10 observed no association. Ritz et al. (2006) linked birth and death certificates for infants
11 who died between 1989 and 2000 to evaluate the influence of outdoor air pollution on
12 infant death in the South Coast Air Basin of California. The authors examined short- and
13 long-term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case
14 subject's death and reported no association between ambient levels of O3 and neonatal
15 mortality. Hajat et al. (2007) analyzed time-series data of daily infant mortality counts in
16 10 major cities in the UK to quantify any associations with short-term changes in air
17 pollution. When the results from the 10 cities were combined there was no relationship
18 between O3 and neonatal mortality, even after restricting the analysis to just the summer
19 months. Lin et al. (2004a) assessed the impact of short-term changes in air pollutants on
20 the number of daily neonatal deaths in Sao Paulo, Brazil. The authors observed no
21 association between ambient levels of O3 and neonatal mortality.
7.4.10.4 Postneonatal Mortality, 1 Month to 1 Year
22 A number of studies focused on the postneonatal period when examining the effects of
23 O3 on infant mortality. Ritz et al. (2006) linked birth and death certificates for infants
24 who died between 1989 and 2000 to evaluate the influence of outdoor air pollution on
25 infant death in the South Coast Air Basin of California. The authors examined short- and
26 long-term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case
27 subject's death and reported no association between ambient levels of O3 and
28 postneonatal mortality. Woodruff et al. (2008) evaluated the county-level relationship
29 between cause-specific postneonatal infant mortality and long-term early-life exposure
30 (first 2 months of life) to air pollutants across the U.S. Similarly, they found no
31 association between O3 exposure and deaths from respiratory causes. In the U.K., Hajat
32 et al. (2007) analyzed time-series data of daily infant mortality counts in 10 major cities
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1 to quantify any associations with short-term changes in air pollution. When the results
2 from the 10 cities were combined there was no relationship between O3 and postneonatal
3 mortality, even after restricting the analysis to just the summer months. In Ciudad Juarez,
4 Mexico, Romieu et al. (2004b) examined the daily number of deaths between 1997 and
5 2001, estimating the modifying effect of SES on the risk of postneonatal mortality.
6 Ambient O3 concentrations were not related to infant mortality overall, or in any of the
7 SES groups. In a follow-up study, Carbajal-Arroyo (2011) evaluated the relationship of
8 1-h daily max O3 levels with postneonatal infant mortality in the Mexico City
9 Metropolitan Area between 1997 and 2005. Generally, short-term exposure to O3 was not
10 significantly related to infant mortality. However, upon estimating the modifying effect
11 of SES on the risk of postneonatal mortality, the authors found that O3 was statistically
12 significantly related to respiratory mortality among those with low SES. In a separate
13 analysis, the effect of PM10 was evaluated with O3 level quartiles. PM10 alone was related
14 to a significant increase in all-cause mortality. The magnitude of this effect remained the
15 same when only the days when O3 was in the lowest quartile were included in the
16 analyses. However, when only the days when O3 was in the highest quartile were
17 included in the analyses, the magnitude of the PM10 effect increased dramatically
18 (OR=1.06 [95% CI: 0.909, 1.241] for PM10 on days with O3 in lowest quartile; OR=1.26
19 [95% CI: 1.08, 1.47] for PM10 on days with O3 in the highest quartile. These results
20 suggest that while O3 alone may not have an effect on infant mortality, it may serve to
21 potentiate the observed effect of PM10 on infant mortality.
22 Tsai et al. (2006) used a case-crossover analysis to examine the relationship between
23 short-term exposure to air pollution and postneonatal mortality in Kaohsiung, Taiwan
24 during the period 1994-2000. The risk of postneonatal deaths was 1.023 (95% CI: 0.564,
25 1.858) per 10-ppb increase in 24-h avg O3. The confidence interval for this effect
26 estimate is very wide, likely due to the small number of infants that died each day,
27 making it difficult to interpret this result. Several other studies conducted in Asia did not
28 find any association between O3 concentrations and infant mortality in the postneonatal
29 period. Ha et al. (2003) conducted a daily time-series study in Seoul, Korea to evaluate
30 the effect of short-term changes in ambient 8-h O3 concentrations on postneonatal
31 mortality. Son et al. (2008) examined the relationship between air pollution and
32 postneonatal mortality from all causes among firstborn infants in Seoul, Korea during
33 1999-2003. Yang et al. (2006) used a case-crossover analysis to examine the relationship
34 between air pollution exposure and postneonatal mortality in Taipei, Taiwan for the
35 period 1994-2000. The authors observed no associations between ambient levels of O3
36 and postneonatal mortality.
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7.4.10.5 Sudden Infant Death Syndrome
1 The strongest evidence for an association between ambient O3 concentrations and SIDS
2 comes from a study that evaluated the county-level relationship between SIDS and long-
3 term early-life exposure (first 2 months of life) to air pollutants across the U.S.(Woodruff
4 et al.. 2008). The authors observed a 1.20 (95% CI: 1.09, 1.32) odds ratio for a 10-ppb
5 increase in O3 and deaths from SIDS. There was a monotonic increase in odds of SIDS
6 for each quartile of O3 exposure compared with the lowest quartile (highest quartile
7 OR=1.51; [95% CI: 1.17, 1.96]). In a multi-pollutant model including PM10 or PM25, CO
8 and SO2, the OR for SIDS and O3 was not substantially lower than that found in the
9 single-pollutant model. When examined by season, the relationship between SIDS deaths
10 and O3 was generally consistent across seasons with a slight increase for those babies
11 born in the summer. When stratified by birth weight, the OR for LEW babies was 1.27
12 (95% CI: 0.95, 1.69) per 10-ppb increase in O3 and the OR for normal weight babies was
13 1.16 (95% CI: 1.01, 1.32) per 10-ppb increase in O3.
14 Conversely, two additional studies reported no association between ambient levels of O3
15 and SIDS. Ritz et al. (2006) linked birth and death certificates for infants who died
16 between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death
17 in the South Coast Air Basin of California. The authors examined short- and long-term
18 exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case subject's
19 death and reported no association between ambient levels of O3 and SIDS. Dales et al.
20 (2004) used time-series analyses to compare the daily mortality rates for SIDS and short-
21 term air pollution concentrations in 12 Canadian cities during the period of 1984-1999.
22 Increased daily rates of SIDS were associated with previous day increases in the levels of
23 SO2, NO2, and CO, but not O3 or PM2 5.
24 Table 7-9 provides a brief overview of the epidemiologic studies of infant mortality.
25 These studies have focused on short-term exposure windows (e.g., 1-3 days) and long-
26 term exposure windows (e.g., up to 6 months). Collectively, they provide no evidence for
27 an association between ambient O3 concentrations and infant mortality.
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Table 7-9 Brief summary of infant mortality studies
Study
Pereiraetal. (1998)
Diaz etal. (2004)
Loomisetal. (1999)
Ritz etal. (2006)
Haiat etal. (2007)
Lin et al. (2004a)
Ha et al. (2003)
Romieu et al. (2004b)
Carbajal-Arroyo et al.
(2011)
Son et al. (2008)
Tsai et al. (2006)
Woodruff etal. (2008)
Yana etal. (2006)
Dales etal. (2004)
Location
Sao Paulo, Brazil
Madrid, Spain
Mexico City, Mexico
Southern California
10 Cities in the UK
Sao Paulo, Brazil
Seoul, South Korea
Ciudad Juarez,
Mexico
Mexico City, Mexico
Seoul, South Korea
Kaohsiung, Taiwan
Nationwide, US
Taipei, Taiwan
12 Canadian cities
Mean Os (ppb)
1-h max: 33.8
24-havg:11.4
24-h avg: 44.1
1-h max: 163.5
24-h avg: 21 .9-22.1
24-h avg: 20.5-42.6
24-h avg: 38.06
8-havg:21.2
8-h avg: 43.43-55.1 2
1-h max: 103.0
8-ha avg: 25.61
24-h avg: 23.60
24-h avg: 26.6
24-h avg: 18.14
24-h: 31 .77
Exposure
Assessment
Citywide avg
Citywide avg
1 monitor
Nearest Monitor
Citywide avg
Citywide avg
Citywide avg
Citywide avg
Citywide avg
Citywide avg
Citywide avg
County wide avg
Citywide avg
Citywide avg
Effect Estimate3 (95% Cl):
10-2:1.00(0.99,1.01)
NR
LO: 0.99 (0.97, 1 .02)
11:0.99(0.96,1.01)
L2: 1 .00 (0.98, 1 .03)
L3: 1 .03 (1 .00, 1 .05)
L4: 1 .01 (0.98, 1 .03)
L5: 1 .02 (0.99, 1 .04)
10-2:1.02(0.99,1.05)
2 wk before death: 1 .03 (0.93, 1 .14)
1 mo before death: NR
2 mo before death: 0.93 (0.89, 0.97)
6 mo before death: NR
10-2:1.00(0.96,1.06)
10:1.00(0.99,1.01)
LO: 0.93 (0.90, 0.96)
L1: 0.96 (0.90, 1.03)
L2: 0.97 (0.91 , 1 .04)
LO-1 cum: 0.96 (0.89, 1.04)
LO-2 cum: 0.94 (0.87, 1.02)
LO: 1 .00 (0.99, 1 .00)
L1: 0.99 (0.99, 0.99)
L2: 0.99 (0.99, 1.00)
LO-2: 0.99 (0.99, 1.00)
L(NR): 0.984 (0.976, 0.992)b
LO-2 cum: 1.02 (0.56, 1.86)
First 2 mo of life: 1.04 (0.98, 1.10)
LO-2 cum:1. 00 (0.62, 1.61)
LO:NR
L1:NR
L2:NR
L3:NR
L4:NR
L5:NR
Multiday lags of 2-6 days: NR
"Relative risk of infant mortality per 10 ppb change in 03
bNo increment provided
LO = Lag 0, L1= Lag 1, L2 = Lag 2, L3 = Lag 3, L4 = Lag 4, L5 = Lag 5, L6 = Lag 6
NR: No quantitative results reported
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Table 7-10 Summary of Key Reproductive and Developmental Toxicological
Studies
Study
Sharkhuu et
al. (2011)
Bignami et al.
(1994)
Haro and Paz
(1993)
Lopez et al.
(2008)
Auten et al.
(2009)
Plopper et al.
(2007)
Fanucchi et
al. (2006)
Dell'Omo et
al. (1995)
Santucci et al.
(2006)
Hanetal.
(2011)
Campos-
Bedolla et al.
(2002)
Kavlocketal.
(1980)
Jedlinska-
Krakowska et
al. (2006)
Model
Pregnant mice;
BALB/c; F; GD9-
18; effects in
offspring
Pregnant CD-1
dams(PD7-17)
Rat dams,
Exposure over
the entirety of
pregnancy;
Rats; Pregnant
dams;GD1-
GD18, GD20, or
GD21.
C57BL/6 mouse
pups
Infant rhesus
monkeys
Infant male
Rhesus
monkeys, post-
natal exposure
CD-1 Mouse
dams and pups
CD-1 Mouse
dams
Rat; Sprague
Dawley, M & F;
PND13
Pregnant Rats;
Sprague Dawley
(GD5, 10 or 18)
CD-1 mice;
(pregnancy day
7-17)
5 month old male
Wistar Hannover
rats
03
(ppm)
0.4, 0.8,
or 1.2
0.4, 0.8
or 1.2
1.0
1.0
1.0
0.5
0.5
0.6
0.3 or
0.6
0.6
3.0
0.4, 0.8
and 1.2
3.0
Exposure Duration
Continuously for 10
consecutive days
Continuous
12h/day during dark cycle
(12h/day, out to either
GD18, GD20orGD21)
3 h/day, every other day,
thrice weekly for 4 weeks
Postnatal, PND30-6month
of age, 5 months of cyclic
exposure, 5days03
followed by 9 days of
filtered air, 8h/day.
5 months of episodic
exposure, age 1 month-
age 6 months, 5 days 03
followed by 9 days of
filtered air, 8h/day.
6 days before breeding to
weaning at PND21
Dam exposure prior to
mating through GD17.
3 h, BALF examined 10h
after 0 3 exposure
1 h on one day of
gestation, uteri collected
16-1 8 h later
Continuous, pregnancy
day 7-1 7
0.5 ppm, 5h/dayfor50
days
Effects
Dams: Decreased number of dams reaching parturition. Offspring: 1-
Decreased birth weights. 2-Decreased rate of postnatal growth (body
weight). 3-lmpaired delayed type hypersensitivity.4-No effect on humoral
immunity. 5-Significantly affected allergic airway inflammation markers
(eosinophilia, IgE) in female offspring sensitized early in life. 6-BALF LDH
significantly elevated in female offspring.
Reproductive success was not affected by 03 exposure (PD7-17,
proportion of successful pregnancies, litter size, ex ratio, frequency of still
birth, or neonatal mortality). Ozone acted as a transient anorexigen in
pregnant dams.
Decreased birth weight and postnatal body weight of offspring out to PND
90. Ozone-exposed pups manifested with inverted sleep-wake patterns or
circadian rhythm phase-shift.
03 induced delayed maturation of near term rodent bronchioles, with
ultra-structural damage to bronchiolar epithelium.
Postnatal 03 exposure significantly increased lung inflammatory cytokine
levels; this was further exacerbated with gestational PM exposure.
Non-significant increases airway resistance and airway responsiveness
with 03 or inhaled allergen alone. Allergen + 03 produced additive
changes in both measures.
Cellular changes and significant structural changes in the distal
respiratory tract in infant rhesus monkeys exposed to 03 postnatally.
Laterality changes in offspring: Ozone exposed pups showed a turning
preference (left turns) distinct from air exposed controls (clockwise turns)
as adults.
Developmental 03 caused increased defensive/submissive behavior in
offspring. 03 exposed offspring also had significant elevations of striatal
BDNF and hippocampal NGF v. air exposed controls.
BALF polymorphonuclear leukocytes and total BALF protein were
significantly elevated in 03 exposed pups. Lung tissue from 03 exposed
pups had significant elevations of manganese superoxide dismutase
(SOD) protein and significant decrements of extra-cellular SOD protein.
Ozone inhalation modifies the contractile response of the pregnant
uterus. The 03 exposed pregnant uteri had significant increases in the
maximum response to acetyl choline stimulation at GD5 and 10; they also
had a significant increase in maximal response to oxytocin at GD 5.
03 induced decrements in postnatal body weight gain. When 03 was co-
administered with sodium salicylate, 03 synergistically increased the rate
of pup resorption (1 .0 ppm GD9-12).
Histopathological evidence of impaired spermatogenesis (round
spermatids/ 21 spermatocytes, giant spermatid cells, and focal epithelial
desquamation with denudation to the 22 basement membrane). Vitamin E
exposure concomitant with 03 protected against pathological changes
but Vitamin C did not.
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7.4.11 Summary and Causal Determination
1 The 2006 O3 AQCD concluded that the limited number of studies that investigated O3
2 demonstrated no associations between O3 and birth outcomes, with the possible
3 exception of birth defects. The current review included an expanded body of evidence on
4 the associations between O3 and reproductive and developmental effects. Recent
5 epidemiologic and toxicological studies provide evidence for an effect of prenatal
6 exposure to O3 on pulmonary structure and function, including lung function changes in
7 the newborn, incident asthma, ultrastructural changes in bronchiole development,
8 alterations in placental and pup cytokines, and increased pup airway hyper-reactivity.
9 Also, there is limited toxicological evidence for an effect of prenatal and early life
10 exposure on central nervous system effects, including laterality, brain morphology,
11 neurobehavioral abnormalities, and sleep aberration. Recent epidemiologic studies have
12 begun to explore the effects of O3 on sperm quality, and provide limited evidence for
13 decrements in sperm concentration, while there is limited toxicological evidence for
14 testicular degeneration associated with O3.
15 While the collective evidence for many of the birth outcomes examined is generally
16 inconsistent (including birth defects), there are several well-designed, well-conducted
17 studies that indicate an association between O3 and adverse outcomes. For example, as
18 part of the southern California Children's Health Study, Salam et al. (2005) observed a
19 concentration-response relationship of decreasing birth weight with increasing O3
20 concentrations averaged over the entire pregnancy that was clearest above the 30-ppb
21 level (see Figure 7-4). Simiarly,Hansen et al. (2008) utilized fetal ultrasonic
22 measurements and found a change in ultrasound measurements associated with O3 during
23 days 31-60 of gestation indicated that increasing O3 concentration decreased an
24 ultrasound measurement for women living within 2 km of the monitoring site.
25 There is no evidence that prenatal or early life O3 concentrations are associated with
26 infant mortality. Collectively, there is limited though positive toxicological evidence for
27 O3-induced developmental effects, including effects on pulmonary structure and function
28 and central nervous system effects. Limited epidemiologic evidence for an effect on
29 prenatal O3 exposure on respiratory development provides coherence with the effects
30 observed in toxicological studies. There is also limited epidemiologic evidence for an
31 association with O3 concentration and decreased sperm concentration. A recent
32 toxicological study provides limited evidence for a possible biological mechanism
33 (histopathology showing impaired spermatogenesis) for such an association.
34 Additionally, though the evidence for an association between O3 concentrations and
35 adverse birth outcomes is generally inconsistent, there are several influential studies that
36 indicate an association with reduced birth weight and restricted fetal growth. Taking into
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1 consideration the positive evidence for developmental and reproductive outcomes from
2 toxicological and epidemiological studies, and the few influential birth outcome studies,
3 the evidence is suggestive of a causal relationship between long-term exposures to O3
4 and reproductive and developmental effects.
7.5 Central Nervous System Effects
7.5.1 Effects on the Brain and Behavior
5 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are
6 associated with alterations in neurotransmitters, motor activity, short and long term
7 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have
8 been observed. Reports of headache, dizziness, and irritation of the nose with O3
9 exposure are common complaints in humans, and some behavioral changes in animals
10 may be related to these symptoms rather than indicative of neurotoxicity. Research in the
11 area of O3-induced neurotoxicity has notably increased over the past few years, and new
12 studies examining the effects of long-term exposure have demonstrated progressive
13 damage in various regions of the brains of rodents in conjunction with altered behavior.
14 Evidence from epidemiologic studies has been more limited. A recently published
15 epidemiologic study examined the association between O3 exposure and neurobehavioral
16 effects. Chen et al. (2009) utilized data from the NHANES III cohort to study the
17 relationship between O3 levels (mean annual O3 concentration 26.5 ppb) and
18 neurobehavioral effects among adults aged 20-59 years. The authors observed an
19 association between annual exposure to O3 and tests measuring coding ability (symbol-
20 digit substitution test) and attention/short-term memory (serial-digit learning test). Each
21 10-ppb increase in annual O3 levels corresponded to an aging-related cognitive
22 performance decline of 3.5 yr for coding ability and 5.3 years for attention/short-term
23 memory. These associations persisted in both crude and adjusted models. There was no
24 association between O3 levels and reaction time tests. The authors concluded that overall,
25 there is an association between long-term O3 exposure and reduced performance on
26 neurobehavioral tests.
27 A number of new toxicological studies demonstrate various perturbations in neurologic
28 function or histology with long-term exposure to O3, including changes similar to those
29 observed in neurodegenerative disorders such as Parkinson's and Alzheimer's disease
30 pathologies in relevant regions of the brain (Table 7-11). The central nervous system is
31 very sensitive to oxidative stress, due in part to its high content of polyunsaturated fatty
32 acids, high rate of oxygen consumption, and low antioxidant enzyme capacity. Oxidative
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1 stress has been identified as one of the pathophysiological mechanisms underlying
2 neurodegenerative disease (Simonian and Coyle. 1996). and it is believed to play a role in
3 altering hippocampal function, which causes cognitive deficits with aging (Vanguilder
4 and Freeman. 2011). A particularly common finding in studies of O3-exposed rats is lipid
5 peroxidation in the brain, especially in the hippocampus, which is important for higher
6 cognitive function including contextual memory acquisition. Performance in passive
7 avoidance learning tests is impaired when the hippocampus is injured. For example, in a
8 subchronic study, exposure of rats to 0.25 ppm O3 (4 h/day) for 15-90 days caused a
9 complex array of responses, including a time-dependent increase in lipid peroxidation
10 products and immunohistochemical changes in the hippocampus that were correlated
11 with decrements in passive avoidance behavioral tests (Rivas-Arancibia et al., 2010).
12 Changes included increased numbers of activated microglia, a sign of inflammation, and
13 progressive neurodegeneration. Notably, continued exposure tends to bring about
14 progressive, cumulative damage, as shown by this study (Rivas-Arancibia et al.. 2010)
15 and others (Santiago-Lopez et al., 2010; Guevara-Guzman et al., 2009; Angoa-Perez et
16 al.. 2006). The effects of O3 on passive avoidance test performance were particularly
17 evident at 90 days for both short- and long-term memory. The greatest extent of cell loss
18 was also observed at this time point, whereas lipid peroxidation did not increase much
19 beyond 60 days of exposure.
20 The substantia nigra is another region of the brain affected by O3, and seems particularly
21 sensitive to oxidative stress because the metabolism of dopamine, central to its function,
22 is an oxidative process perturbed by redox imbalance. Oxidative stress has been
23 implicated in the premature death of substantia nigra dopamine neurons in Parkinson's
24 disease. Progressive damage has been found in the substantia nigra of male rats after 15,
25 30, and 60 days of exposure to 0.25 ppm O3 for 4 h/day. Santiago-Lopez and colleagues
26 (2010) observed a reduction dopaminergic neurons within the substantia nigra over time,
27 with a complete loss of normal morphology in the remaining cells and virtually no
28 dopamine immunoreactivity at 60 days. This was accompanied by an increase in p53
29 levels and nuclear translocation, a process associated with programmed cell death.
30 Similarly, Angoa-Perez et al. (2006) have shown progressive lipoperoxidation in the
31 substantia nigra and a decrease in nigral neurons in ovariectomized female rats exposed
32 to 0.25 ppm O3, 4h/day, for 7-60 days. Lipid peroxidation effectively doubled between
33 the 30 and 60 day time points. Total nigral cell number was also diminished to the
34 greatest extent at 60 days, and cell loss was particularly evident in the tyrosine
35 hydroxylase positive cell population (90%), indicating a selective loss of dopamine
36 neurons or a loss of dopamine pathway functionality.
37 The olfactory bulb also undergoes oxidative damage in O3-exposed animals, in some
3 8 cases altering olfactory-dependent behavior. Lipid peroxidation was observed in the
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1
2
3
4
5
6
7
9
10
11
12
13
14
15
16
olfactory bulbs of ovariectomized female rats exposed to 0.25 ppm O3 (4 h/day) for 30 or
60 days (Guevara-Guzman et al.. 2009). O3 also induced decrements in a selective
olfactory recognition memory test, which were significantly greater at 60 days compared
to 30 days, and the authors note that early deficits in odor perception and memory are
components of human neurodegenerative diseases. The decrements in olfactory memory
did not appear to be due to damaged olfactory perception based on other tests early on,
but by 60 days deficits in olfactory perception had emerged.
Memory deficits and associated morphological changes can be attenuated by
administration of a-tocopherol (Guerrero etal.. 1999). taurine (Rivas-Arancibia et al..
2000). and estradiol (Guevara-Guzman et al.. 2009; Angoa-Perez et al.. 2006). all of
which have antioxidant properties. In the study by Angoa-Perez et al. (2006) described
above, estradiol seemed particularly effective at protecting against lipid peroxidation and
nigral cell loss at 60 days compared to shorter exposure durations. The same was true for
amelioration of decrements in olfactory recognition memory (Guevara-Guzman et al..
2009). although protection against lipid peroxidation was similar for the 30 and 60 day
exposures.
Table 7-11
Study
Angoa-Perez et al.
(2006)
Central nervous system effects of long-term Os exposure in rats
Model
Rat; Wistar; F;
Weight: 300 g;
Ovariectomized
O3 (ppm) Exposure Duration
0.25 7 to 60 days, 4 h/day,
5 days/wk
Effects
Long-term estradiol treatment protected against
03-induced oxidative damage to nigral
dopamine neurons, lipid peroxidation, and loss
of tyrosine hydrolase-immunopositive cells.
Guevara-Guzman et Rat; Wistar; F;
al. (2009) Weight: 264 g;
Ovariectomized
0.25
30 and 60 days, 4h/day
Long-term estradiol treatment protected against
03-induced oxidative stress and decreases in a
and p estrogen receptors and dopamine p-
hydroxlyase in olfactory bulb, and deficits in
olfactory social recognition memory and
chocolate recognition.
Rivas-Arancibia et
al. (2010)
Rat; Wistar; M;
Weight: 250-300 g
0.25
15 to 90 days, 4h/day
Ozone produced significant increases in lipid
peroxidation in the hippocampus, and altered
the number of p53 positive immunoreactive
cells, activated and phagocytic microglia, GFAP
immunoreactive cells, double cortine cells, and
short- and long-term memory-retention latency
Santiago-Lopez et
al. (2010)
Santucci et al.
(2006)
Rat; Wistar; M; 0.25
Weight: 250-300 g
Mice;CD-1;M; 0.3; 0.6
18 weeks old
15, 30, and 60 days,
4 h/day
Females continuously
exposed from 30 days prior
to breeding untilGD17
Progressive loss of dopamine reactivity in the
substantia nigra, along with morphological
changes. Increased p53 levels and nuclear
translocation.
Upon behavioral challenge with another male,
there was a significant increase in defensive
and freezing postures and decrease in the
frequency of nose-sniffing. These behavioral
changes were accompanied by a significant
increase in BDNF in the striatum and a
decrease of NGF in the hippocampus.
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1 CNS effects have also been demonstrated in adult mice whose only exposure to O3
2 occurred while in utero, a period particularly critical for brain development. Santucci et
3 al. (2006) investigated behavioral effects and gene expression after in utero exposure of
4 mice to 0.3 or 0.6 ppm O3. Exposure began 30 days prior to mating and continued
5 throughout gestation. Testing of adult animals demonstrated increased
6 defensive/submissive behavior and reduced social investigation were observed in both the
7 0.3 and 0.6 ppm O3 groups. Changes in gene expression of brain-derived neurotrophic
8 factor (BDNF, increased in striatum) and nerve growth factor (NGF, decreased in
9 hippocampus) accompanied these behavioral changes. BDNF and NGF are involved in
10 neuronal organization and the growth, maintenance, and survival of neurons during early
11 development and in adulthood. This study and two others using short-term exposures
12 demonstrate that CNS effects can occur as a result of in utero exposure to O3, and
13 although the mode of action of these effects is not known, it has been suggested that
14 circulating lipid peroxidation products may play a role (Boussouar et al.. 2009).
15 Importantly, these CNS effects occurred in rodent models after in utero only exposure to
16 (semi-) relevant concentrations of O 3.
7.5.2 Summary and Causal Determination
17 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are
18 associated with alterations in neurotransmitters, motor activity, short and long term
19 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have
20 been observed. However, evidence regarding chronic exposure and neurobehavioral
21 effects was not available. Recent research in the area of O3-induced neurotoxicity has
22 included several long-term exposure studies. Notably, the first epidemiologic study to
23 examine the relationship between O3 exposure and neurobehavioral effects observed an
24 association between annual O3 levels and an aging-related cognitive performance decline
25 in tests measuring coding ability and attention/short-term memory. This observation is
26 supported by studies in rodents which demonstrate progressive oxidative stress and
27 damage in the brain and associated decrements in behavioral tests, including those
28 measuring memory, after subchronic exposure to 0.25 ppm O3. Additionally,
29 neurobehavioral changes are evident in animals whose only exposure to O3 occurred in
30 utero. Collectively, the limited epidemiologic and toxicological evidence is coherent and
31 suggestive of a causal relationship between O3 exposure and CNS effects.
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7.6 Carcinogenic and Genotoxic Potential of Ozone
7.6.1 Introduction
1 The radiomimetic and clastogenic qualities of O3, combined with its ability to stimulate
2 proliferation of cells in the respiratory tract, have suggested that O3 could act as a
3 carcinogen. However, toxicological studies of tumorigenesis in the rodent lung have
4 yielded mixed and often confusing results, and the epidemiologic evidence is equally
5 conflicted. The 2006 O3 AQCD concluded that, "the weight of evidence from recent
6 animal toxicological studies and a very limited number of epidemiologic studies do not
7 support ambient O3 as a pulmonary carcinogen" 2(U.S. EPA, 2006b).
8 Multiple epidemiologic studies reported in the 2006 O3 AQCD examined the association
9 between O3 exposure and cancer. The largest of these studies, by Pope et al. (2002).
10 included 500,000 adults from the American Cancer Society's (ACS) Cancer Prevention II
11 study. In this study, no association was observed between O3 and lung cancer mortality.
12 The Adventist Health Study of Smog (AHSMOG) also examined the association between
13 O3 and lung cancer mortality (Abbey etal.. 1999). There was a positive association
14 between O3 levels and lung cancer mortality among men. No association was reported for
15 women. Another study using the AHSMOG cohort assessed the risk of incident lung
16 cancer (Beeson et al., 1998). Among males, an association with incidence of lung cancer
17 was observed with increasing O3 concentrations. When stratified by smoking status, the
18 association persisted among never smokers but was null for former smokers. No
19 association was detected for females. The Six Cities Study examined various air
20 pollutants and mortality but did not specifically explore the association between O3
21 concentrations and lung cancer mortality due to low variability in O3 levels across the
22 cities (Dockery et al., 1993). An ecologic study performed in Sao Paulo City, Brazil
23 examined the correlations between O3 levels in four of the city districts and incident
24 cancer of the larynx and lung reported in 1997 (Pereira et al., 2005). A correlation
25 between the average number of days O3 levels exceeded air quality standards from 1981
26 to 1990 and cancer incidence was present for larynx cancer but not for lung cancer.
27 Early toxicological research demonstrated lung adenoma3 acceleration in mice with daily
28 exposure to 1 ppm over 15 months (Stokinger. 1962). Later work demonstrated a
29 significant increase in lung tumor numbers in one strain of mouse (A/J) but not another
2 The toxicological evidence is presented in detail in Table 6-18 on p. 6-116 of the 1996 O3 AQCD and Table AX5-13 on p.AX5-43
of the 2006 O3 AQCD.
3 NOTE: Although adenomas are benign, over time they may progress to become malignant, at which point they are called
adenocarcinomas. Adenocarcinoma is the predominant lung cancer subtype in most countries, and is the only lung cancer found in
nonsmokers. From page 8-33 of the 1970 O3 AQCD: "No true lung cancers have been reported, however, from experimental
exposures to either O3 alone or any other combination or ingredient of photochemical oxidants."
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1 after exposure to 0.3-0.8 ppm O3 (Lastetal.. 1987; Hassett et al.. 1985). The A/J mouse
2 strain is known to have a high incidence of spontaneous adenomas, and further studies
3 using this strain found a statistically significant increase in lung tumor incidence after a
4 9-month exposure to 0.5 ppm and incidence and multiplicity after a 5 month exposure to
5 0.12 ppm with a 4-month recovery period (Witschi et al.. 1999). However, these findings
6 were discounted by the study authors due to the lack of a clear dose response, and results
7 from the Hassett et al. 1985 and Last et al. 1987 studies were retrospectively deemed
8 spurious based on what appeared to be unusually low spontaneous tumor incidences in
9 the control groups (Witschi. 1991). A study of carcinogenicity of O3 by the National
10 Toxicology Program (NTP. 1994) reported increased incidences of alveolar/bronchiolar
11 adenoma or carcinoma (combined) in female B6C3FJ mice exposed over 2 years to
12 1.0 ppm O3, but not 0.12 or .5 ppm. No effect was detected in male mice. For a lifetime
13 exposure to 0.5 or 1.0 ppm O3, an increase in the number of female mice with adenomas
14 (but not carcinomas or total neoplasms) was found. The number of total neoplasms was
15 also unaffected in male mice, but there was a marginally increased incidence of
16 carcinoma in males exposed to 0.5 and 1.0 ppm. Thus there was equivocal evidence of
17 carcinogenic activity in male mice and some evidence of carcinogenic activity of O3 in
18 females. Some semblance of a dose-response relationship was also evident in this study.
19 Experimental details of the NTP study are available in Table 6-19 on p. 6-121 of the 1996
20 O3 AQCD.
21 In Fischer-344/N rats (50 of each sex per group), neither a 2-year nor lifetime exposure to
22 O3 ranging from 0.12 to 1.0 ppm was found to be carcinogenic (Boorman et al.. 1994).
23 However, a marginally significant carcinogenic effect of 0.2 ppm O3 was reported in a
24 study of male Sprague-Dawley rats exposed for 6 months (n = 50) (Monchaux et al..
25 1996). These two studies also examined co-carcinogenicity of O3 with NNK4 (Boorman
26 etal.. 1994) or a relatively high dose of radon (Monchaux et al.. 1996). finding no
27 enhancement of NNK related tumors and a slight non-significant increase in tumor
28 incidence after combined exposure with radon, respectively. Another study exploring co-
29 carcinogenicity was conducted in hamsters. Not only was there no enhancement of
30 chemically induced tumors in the peripheral lung or nasal cavity, but results suggested
31 that O3 could potentially delay or inhibit tumor development (Witschi et al.. 1993). Thus
32 there is no concrete evidence that O3 can act as a co-carcinogen.
33 Immune surveillance is an important defense against cancer, and it should be noted that
34 natural killer (NK) cells, which destroy tumor cells in the lung, appear to be inhibited by
35 higher doses of O3 and either unaffected or stimulated at lower doses (Section 6.2.5.4,
4 4-(N-nitrosomethylamino)-1 -(3-pyridyl)-1 -butanone
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1 Infection and Adaptive Immunity). This aspect of tumorigenesis adds yet another layer of
2 complexity which may be reflected by conflicting results across studies.
3 The following sections will examine epidemiologic studies of cancer incidence and
4 mortality and toxicological studies that have been published since the 2006 O3 AQCD.
5 An epidemiologic study has been published with cancer as the outcome; most
6 epidemiologic studies examine markers of exposure or susceptibility.
7.6.2 Lung Cancer Incidence and Mortality
7 A recent re-analysis of the full ACS CPSII cohort by the Health Effects Institute is the
8 only epidemiologic study that has explored the association between O3 and cancer
9 mortality since the last O3 AQCD. Krewski et al. (2009) conducted an extended follow-
10 up of the cohort (1982-2000). Mean O3 levels [obtained from the Aerometric Information
11 Retrieval System (AIRS) for 1980] were 22.91 ppb for the full year and 30.15 ppb for the
12 summer months (April-September). No association was reported between lung cancer
13 mortality and O3 (HR=1.00 [95% CI: 0.96-1.04] per 10 ppb O3). Additionally, no
14 association was observed when O3 was restricted to the summer months. There was also
15 no association present in a sub-analysis of the cohort examining the relationship between
16 O3 and lung cancer mortality in the Los Angeles area.
17 Since the 2006 O3 AQCD, two toxicological studies have examined potential
18 carcinogenicity of O3 (Kim and Cho. 2009a. b). Looking across both studies, which used
19 the same mouse strain as the National Toxicology Program study described above (NTP.
20 1994). 0.5 ppm O3 alone or in conjunction with chemical tumor inducers did not enhance
21 lung tumor incidence in males or females. However, a 10% incidence of oviductal
22 carcinoma was observed in mice exposed to 0.5 ppm O3 for 16 weeks. The implications
23 of this observation are unclear, particularly in light of the lack of statistical information
24 reported. Additionally, there is no mention of oviductal carcinoma after 32 weeks of
25 exposure, and no oviductal carcinoma was observed after one year of exposure. The NTP
26 study did not report any increase in tumors at extrapulmonary sites.
7.6.3 DMA Damage
27 The potential for genotoxic effects relating to O3 exposure was predicted from the
28 radiomimetic properties of O3. The decomposition of O3 in water produces OH and HO2
29 radicals, the same species that are generally considered to be the biologically active
30 products of ionizing radiation. Ozone has been observed to cause degradation of DNA in
31 a number of different models and bacterial strains. The toxic effects of O3 have been
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1 generally assumed to be confined to the tissues directly in contact with the gas, such as
2 the respiratory epithelium. Due to the highly reactive nature of O3, little systemic
3 absorption is predicted. Zelac et al. (1971a. b), however, reported a significant increase in
4 chromosome aberrations in peripheral blood lymphocytes from Chinese hamsters
5 exposed to 0.2 ppm for 5 hours. Other in vivo exposure studies found increased DNA
6 strand breaks in respiratory cells from guinea pigs (Ferng et al.. 1997) and mice
7 (Bornholdt et al., 2002) but only with exposure to higher doses of O3 (1 ppm for 72 hours
8 and 1 or 2 ppm for 90 minutes, respectively). In other studies there were no observations
9 of chromosomal aberrations in germ cells, but mutagenic effects have been seen in
10 offspring of mice exposed to 0.2 ppm during gestation (blepharophimosis or dysplasia of
11 the eyelids). The overall evidence for mutagenic activity from in vitro studies is positive,
12 and in the National Toxicology Program report described above, O3 was found to be
13 mutagenic in Salmonella, with and without S9 metabolic activation. No new
14 toxicological studies of DNA damage have become available since the 2006 O3 AQCD.
15 A number of epidemiologic studies looked at the association between O3 and DNA and
16 cellular level damages. These changes may be relevant to mechanisms leading to cancers
17 development and serve as early indicators of elevated risk of mutagenicity.
18 Two studies performed in California examined cytogenetic damage in relation to O3
19 exposures. Huen et al. (2006) examined cytogenetic damage among African American
20 children and their mothers in Oakland, CA. Increased O3 (mean monthly 8-h O3
21 concentrations ranged from about 30 ppb in April to 14 ppb in November) was associated
22 with increased cytogenetic damage (micronuclei frequency among lymphocytes and
23 buccal cells) even after adjustment for household/personal smoking status and distance-
24 weighted traffic density. Chen et al. (2006a) recruited college students at the University
25 or California, Berkeley who reported never smoking and compared their levels of
26 cytogenetic damage (micronuclei frequency from buccal cells) in the spring and fall.
27 Cytogenetic damage was greater in the fall, which the authors attributed to the increase in
28 O3 over the summer. However, O3 levels over 2, 7, 10, 14, or 30 days (concentrations not
29 given) before collection of buccal cells did not correlate with cytogenetic damage.
30 Estimated lifetime O3 exposure was also not correlated with cytogenetic damage.
31 Additionally, the authors exposed a subset of the students (n=15) to 200 ppb O3 for
32 4 hours while the students exercised intermittently. Ozone was found to be associated
33 with an increase in cytogenetic damage in degenerated cells but not in normal cells 9-
34 10 days after exposure. Increased cytogenetic damage was also noted in peripheral blood
35 lymphocytes collected 18 hours after exposure.
36 A study performed in Mexico recruited 55 male workers working indoors (n=27) or
37 outdoors (n=28) in Mexico City or Puebla, Mexico in order to study the relationship
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1 between O3 and DNA damage (detected from peripheral blood samples using the Comet
2 assay) (Tovalin et al.. 2006). The median estimated daily O3 concentrations were
3 estimated to be 28.5 ppb for outdoor workers and 5.1 ppb for indoor workers in
4 Mexico City and 36.1 ppb for outdoor workers and 19.5 ppb for indoor workers in
5 Puebla. Overall, a positive correlation between O3 levels and DNA damage was
6 observed. However, when examining the relationship by city and workplace, only DNA
7 damage in outdoor workers in Mexico City remained correlated with O3 levels.
8 Three studies examining the relationship between O3 and DNA-level damage have been
9 performed in Europe. The largest of these was the GenAir case-control study, which was
10 nested within the European Prospective Investigation into Cancer and Nutrition (EPIC)
11 study, and included individuals recruited between 1993 and 1998 from ten European
12 countries. Only non-smokers (must not have smoked for at least 10 years prior to
13 enrollment) were enrolled in the study. The researchers examined DNA adduct levels
14 (DNA bonded to cancer-causing chemicals) and their relationship with O3 concentrations
15 (concentrations not given) (Peluso M Hainaut et al.. 2005). A positive association was
16 seen between DNA adduct levels and O3 concentrations from 1990-1994 but not O3
17 levels from 1995-1999. In adjusted analyses with DNA adduct levels dichotomized as
18 high and low (detectable versus non-detectable), the OR was 1.97 (95% CI: 1.08, 3.58)
19 when comparing the upper tertile of O3 concentration to the lower two tertiles. Two other
20 European studies were conducted in Florence, Italy. The most recent of these enrolled
21 individuals from the EPIC study into a separate study between March and September of
22 1999 (Palli et al.. 2009). The purpose of the study was to examine oxidative DNA
23 damage (determined by Comet assay using blood lymphocytes) in association with
24 varying periods of O3 exposure. The researchers observed that longer periods of high O3
25 exposure (concentrations not given) were more strongly correlated with oxidative DNA
26 damage than shorter exposures (i.e., the rho [p-value] was 0.26 [0.03] for 0-10 days and
27 0.35 [0.002] for 0-90 days). This correlation was stronger among men compared to
28 women. The correlations for all time periods had p-values <0.05 for ex- and never-
29 smokers. For current smokers, the correlation was only observed among time periods <
30 25 days. When adjusted for age, gender, smoking history, traffic pollution exposure,
31 period of blood draw, and area of residence, the association between O3 levels and
32 oxidative DNA damage was positive for O3 levels 0-60 days, 0-75 days, and 0-90 days
33 prior to blood draw. Positive, statistically significant associations were not observed
34 among shorter time periods. The other study performed in Florence recruited healthy
35 volunteers who reported being non-smokers or light smokers (Giovannelli et al.. 2006).
36 The estimated O3 levels during the study ranged from approximately 4-40 ppb for 3-day
37 avgs, 5-35 ppb for 7-day avgs, and 7.5-32.5 ppb for 30-day avgs. Ozone concentrations
38 were correlated with DNA strand breaks (measured from blood lymphocytes) over longer
39 exposure periods (p-value: 0.002 at 30 days, p-value: 0.04 at 7 days; p-value: 0.17 at
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1 3 days). This association was robust to control for temperature, solar radiation, gender,
2 and age. No association was seen between O3 concentrations and measures of oxidative
3 DNA damage at 3, 7, or 30 days.
7.6.4 Summary and Causal Determination
4 The 2006 O3 AQCD reported that evidence did not support ambient O3 as a pulmonary
5 carcinogen. Since the 2006 O3 AQCD, very few epidemiologic and toxicological studies
6 have been published that examine O3 as a carcinogen, but collectively, study results
7 indicate that O3 may contribute to DNA damage. Overall, the evidence is inadequate to
8 determine if a causal relationship exists between ambient O3 exposures and
9 cancer.
7.7 Mortality
10 A limited number of epidemiologic studies have assessed the relationship between long-
11 term exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an
12 insufficient amount of evidence existed "to suggest a causal relationship between chronic
13 O3 exposure and increased risk for mortality in humans" (U.S. EPA. 2006b). In addition
14 to the infant mortality studies discussed in Section 7.4.9, additional studies have been
15 conducted among adults since the last review; an ecologic study that finds no association
16 between mortality and O3, several reanalyses of the ACS cohort, one of which
17 specifically points to a relationship between long-term O3 exposure and an increased risk
18 of respiratory mortality, and a study of four cohorts of persons with potentially
19 predisposing conditions. These studies supplement the evidence from long-term cohort
20 studies characterized in previous reviews of O3 and are summarized here briefly.
21 In the Harvard Six Cities Study (Dockery et al.. 1993), adjusted mortality rate ratios were
22 examined in relation to long-term mean O3 concentrations in six cities: Topeka, KS; St.
23 Louis, MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mean
24 O3 concentrations from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in
25 Portage. Long-term mean O3 concentrations were not found to be associated with
26 mortality in the six cities. However, the authors noted that "The small differences in O3
27 levels among the (six) cities limited the power of the study to detect associations between
28 mortality and O3 levels." In addition, while total and cardio-pulmonary mortality were
29 considered in this study, respiratory mortality was not specifically considered.
30 In a subsequent large prospective cohort study of approximately 500,000 U.S. adults,
31 Pope et al. (2002) examined the effects of long-term exposure to air pollutants on
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1
2
3
4
5
6
7
8
9
10
1 1
mortality (American Cancer Society, Cancer Prevention Study II). All-cause,
cardiopulmonary, lung cancer and other mortality risk estimates for long-term O3
exposure are shown in Figure 7-5. While consistently positive associations were not
observed between O3 and mortality (effect estimates labeled A in Figure 7-5) , the
mortality risk estimates were larger in magnitude when analyses considered more
accurate exposure metrics, increasing when the entire period was considered (effect
estimates labeled B in Figure 7-5) and becoming marginally significant when the
exposure estimate was restricted to the summer months (July to September; effect
estimates labeled C in Figure 7-5), especially when considering cardiopulmonary deaths.
In contrast, consistent positive and significant effects of PM2 5 were observed for both
lung cancer and cardio-pulmonary mortality.
All Cause Cardiopulmonary
Mortality Mortality
^ ' ;• [ ] [ — 1
^ > i i i „ . a
§ r J
i ' . , i o ° | ? 1
1 (
ABC A B C
., . „ . „ „ .. Number of
Years of Data Collection .. . ... .
Metropolitan Areas
A 1980-1981 134
B 1982-1998 119
C 1982-1998 (July -Sept) 134
I ,ung Cancer
Mortality
I , I
-I ' o
1 °
A B C
Number of Participants
(in thousands)
559
525
557
All Other Ci
Mortalit
I
I f\
\J
A B
1-h Max Os Mean
47.9(11.0)
45.5 (7.3)
59.7(12.8)
a uses
V
I
0
C
(SD)
Source: Reprinted with permission of American Medical Association, Pope et al. (2002).
Figure 7-5 Adjusted ozone-mortality relative risk estimates (95% Cl) by time
period of analysis per subject-weighted mean Os concentration in
the Cancer Prevention Study II by the American Cancer Society.
12
13
14
15
16
A study by Abbey et al. (1999) examined the effects of long-term air pollution exposure,
including O3, on all-cause (n = 1,575), cardiopulmonary (n = 1,029), nonmalignant
respiratory (n = 410), and lung cancer (n = 30) mortality in the long-term prospective
Adventist Health Study of Smog (AHSMOG) of 6,338 nonsmoking, non-Hispanic white
individuals living in California. A particular strength of this study was the extensive
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1 effort devoted to assessing long-term air pollution exposures, including interpolation to
2 residential and work locations from monitoring sites over time and space. No associations
3 with long-term O3 exposure were observed for all cause, cardiopulmonary, and
4 nonmalignant respiratory mortality. In a follow-up, Chen et al. (2005) utilized data from
5 the AHSMOG study and reported no evidence of associations between long-term O3
6 exposure (mean O3 concentration 26.2 ppb) and fatal coronary heart disease. Thus, no
7 association of chronic O3 exposure with mortality outcomes has been detected in this
8 study.
9 Lipfert et al. (2003. 2000) reported positive effects on all-cause mortality for peak O3
10 exposures (95th percentile levels) in the U.S. Veterans Cohort study of approximately
11 50,000 middle-aged men recruited with a diagnosis of hypertension. The actual analysis
12 involved smaller subcohorts based on exposure and mortality follow-up periods. Four
13 separate exposure periods were associated with three mortality follow-up periods. For
14 concurrent exposure periods, peak O3 was positively associated with all-cause mortality,
15 with a 9.4% (95% CI: 0.4, 18.4) excess risk per mean 95th percentile O3 less estimated
16 background level (not stated). "Peak" refers, in this case, to the 95th percentile of
17 the hourly measurements, averaged by year and county. In a further analysis, Lipfert et al.
18 (2003) reported the strongest positive association for concurrent exposure to peak O3 for
19 the subset of subjects with low diastolic blood pressure during the 1982 to 1988 period.
20 Two more recent studies of this cohort focused specifically on traffic density (Lipfert et
21 al.. 2006a; 2006b). Lipfert (2006b) concluded that: "Traffic density is seen to be a
22 significant and robust predictor of survival in this cohort, more so than ambient air
23 quality, with the possible exception of O3," reporting a significant O3 effect even with
24 traffic density included in the model: RR=1.080 per 40 ppb peak O3 (95% CI: 1.019,
25 1.146). However, in Lipfert (2006a), which considers only the EPA Speciation Trends
26 Network (STN) sites, O3 drops to non-significant predictor of total mortality for this
27 cohort. The authors acknowledge that: "Peak O3 has been important in analyses of this
28 cohort for previous periods, but in the STN data set, this variable has limited range and
29 somewhat lower values and its small coefficient of variation results in a relatively large
30 standard error." The restriction to subjects near STN sites likely reduced the power of this
31 analysis, though the size of the remaining subjects considered was not reported in this
32 paper. In addition, these various Veterans Cohort studies considered only total mortality,
33 and did not consider mortality on a by-cause basis.
34 An ecological study in Brisbane, Australia used a geospatial approach to analyze the
35 association of long-term exposure to gaseous air pollution with cardio-respiratory
36 mortality, in the period 1996-2004 (Wang et al.. 2009c). A generalized estimating
37 equations model was employed to investigate the impact of NO2, O3 and SO2, but PM
38 was not addressed. The results indicated that long-term exposure to O3 was not
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1 associated with cardio-respiratory mortality, but the fact that this study considered only
2 one city, and that the range of O3 exposure across that city (23.7-35.6 ppb) was low and
3 slight in variation in comparison to the range of other pollutants across the city, limited
4 study power. In addition, confounding factors (e.g., smoking) could not be addressed at
5 the individual level in this ecological study. Respiratory mortality was not evaluated
6 separately.
7 A recent study by Zanobetti and Schwartz (In Press) examined whether year-to-year
8 variations in 8-h mean daily O3 concentrations for the summer (May-September) around
9 their city-specific long-term trend were associated with year-to-year variations in
10 mortality around its long-term trend. This association was examined among Medicare
11 participants with potentially predisposing conditions, including COPD, diabetes, CHF,
12 and MI, defined as patients discharged alive after an emergency admission for one of
13 these four conditions. The analyses was repeated in 105 cities using available data from
14 1985 through 2006, and the results were combined using methods previously employed
15 by these authors (Zanobetti et al.. 2008; Zanobetti and Schwartz. 2007). This study
16 design eliminated potential confounding by factors that vary across city, which is a
17 common concern in most air pollution cohort studies, and also avoided both confounding
18 by cross-sectional factors that vary by city and the short-term factors that confound daily
19 time-series studies, but are not present in annual analyses. The average 8-h mean daily
20 summer O3 concentrations ranged from 15.6 ppb (Honolulu, HI) to 71.4 ppb
21 (Bakersfield, CA) for the 105 cities. The authors observed associations between yearly
22 fluctuations in summer O3 concentrations and mortality in each of the four cohorts; the
23 hazard ratios (per 10 ppb increment) were 1.12 (95% CI: 1.06, 1.17) for the CHF cohort,
24 1.19 (95% CI 1.12, 1.25) for the MI cohort, 1.14 (95% CI: 1.10, 1.21) forthe diabetes
25 cohort, and 1.14 (95% CI: 1.08, 1.19) for the COPD cohort. A key advantage to this study
26 is that fluctuations from summer to summer in O3 concentrations around long-term level
27 and trend in a specific city are unlikely to be correlated with most other predicators of
28 mortality risk, except for temperature, which was controlled for in the regression. Key
29 limitations of the study were the inability to control for PM2 5, since it was not reliably
30 measured in these cities until 1999, and the inability to separate specific causes of death
31 (e.g., respiratory, cardiovascular), since Medicare does not provide the underlying cause
32 of death.
33 In the most recent follow-up analyses of the ACS cohort (Jerrett et al.. 2009; Smith et al..
34 2009a), the effects of long-term exposure to O3 were evaluated alone, as well as in
35 copollutant models with PM2 5 and components of PM2 5. Jerrett et al. (2009) utilized the
36 ACS cohort with data from 1977 through 2000 (mean O3 concentration ranged from 33.3
37 to 104.0 ppb) and subdivided cardiopulmonary deaths into respiratory and cardiovascular,
38 separately, as opposed to combined into one category, as was done by Pope et al. (2002).
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1 Increases in exposure to O3 were associated with an elevated risk of death from
2 cardiopulmonary, cardiovascular, ischemic heart disease, and respiratory causes.
3 Inclusion of PM2 5 concentrations measured in 1999-2000 as a copollutant attenuated the
4 association with O3 for all end points except death from respiratory causes, for which a
5 significant association persisted (Table 7-12). The association between increased O3
6 concentrations and increased risk of death from respiratory causes was insensitive to the
7 use of a random-effects survival model allowing for spatial clustering within the
8 metropolitan area and state of residence, and adjustment for several ecologic variables
9 considered individually. Subgroup analyses showed that temperature and region of
10 country, but not sex, age at enrollment, body-mass index, education, or PM25
11 concentration, modified the effects of O3 on the risk of death from respiratory causes
12 (i.e., risks were higher at higher temperature, and in the Southeast, Southwest, and Upper
13 Midwest). Ozone threshold analyses indicated that the threshold model was not a better
14 fit to the data (p > 0.05) than a linear representation of the overall O3-mortality
15 association. Overall, this new analysis indicates that long-term exposure to PM2 5
16 increases risk of cardiac death, while long-term exposure to O3 is specifically associated
17 with an increased risk of respiratory death, and suggests that combining cardiovascular
18 and respiratory causes of mortality into one category for analysis may obscure any effect
19 that O3 may have on respiratory-related causes of mortality.
Table 7-12 Relative risk (and 95% Cl) of death attributable to a 10-ppb change
in the ambient Os concentration*
Cause of Death
Any Cause
Cardiopulmonary
Respiratory
Cardiovascular
Ischemic Heart Disease
O3 (96 MSAs)
1 .001 (0.996, 1 .007)
1.014(1.007,1.022)
1.029(1.010,1.048)
1.011 (1.003,1.023)
1.015(1.003,1.026)
O3 (86 MSAs)
1.001 (0.996,1.007)
1.016(1.008,1.024)
1 .027 (1 .007, 1 .046)
1.014(1.005,1.023)
1.017(1.006,1.029)
O3 +PM2.5 (86 MSAs)
0.989 (0.981 , 0.996)
0.992(0.982,1.003)
1.040(1.013,1.067)
0.983 (0.971 , 0.994)
0.973 (0.958, 0.988)
* Ozone concentrations were measured from April to September during the years from 1977 to 2000, with follow-up from 1982 to 2000; changes in
the concentration of PM2.5 of 10 ug per cubic meter were recorded for members of the cohort in 1999 and 2000.
Source: Reprinted with permission of Massachusetts Medical Society (Jerrettetal.. 2009)
20 In a similar analysis, Smith et al. (2009a) used data from 66 MSAs in the ACS cohort to
21 examine the association of O3 concentrations during the warm season and all-cause and
22 cardiopulmonary mortality. Mortality effects were estimated in single pollutant and
23 copollutant models, adjusting for two PM2 5 constituents, sulfate and EC. When all-cause
24 mortality was investigated, there was a 0.8% (95% CI: -0.31, 1.9) increase associated
25 with a 10 ppb increase in O3 concentration. This association was diminished when sulfate
26 or EC were included in the model. There was a 2.48% (95% CI: 0.74, 4.3) increase in
27 cardiopulmonary mortality associated with a 10 ppb increase in O3 concentration. The
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1 cardiopulmonary association was robust to adjustment for sulfate, and diminished, though
2 still positive, after adjustment for EC (1.63% increase; 95% CI: -0.41, 3.7). Smith et al.
3 (2009a) did not specifically separate out cardiovascular and respiratory causes of death
4 from the cardiopulmonary category, as was done by Jerrett et al. (2009).
7.7.1 Summary and Causal Determination
5 The 2006 O3 AQCD concluded that an insufficient amount of evidence existed "to
6 suggest a causal relationship between chronic O3 exposure and increased risk for
7 mortality in humans" (U.S. EPA. 2006b). Several additional studies have been conducted
8 since the last review, including an ecologic study that finds no association between
9 mortality and O3 (Wang et al.. 2009c). a study of four cohorts of Medicare enrollees with
10 potentially predisposing conditions that observes associations between O3 and mortality
11 among each of the cohorts (Zanobetti and Schwartz. In Press), and reanalyses of the ACS
12 cohort that provide weak evidence for an association with cardiopulmonary mortality
13 (Smith et al.. 2009a) and specifically point to a relationship between long-term O3
14 exposure and an increased risk of respiratory mortality (Jerrett et al.. 2009). The findings
15 from the Jerrett et al. (2009) study are consistent and coherent with the evidence from
16 epidemiologic, controlled human exposure, and animal toxicological studies for the
17 effects of short- and long-term exposure to O3 on respiratory effects. Additionally, the
18 evidence for short- and long-term respiratory morbidity provides biological plausibility
19 for mortality due to respiratory disease. Collectively, the evidence is suggestive of a
20 causal relationship between long-term O3 exposures and mortality.
7.8 Overall Summary
21 The evidence reviewed in this chapter describes the recent findings regarding the health
22 effects of long-term exposure to ambient O3 concentrations. Table 7-13 provides an
23 overview of the causal determinations for each of the health categories evaluated.
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Table 7-13 Summary of causal determinations for long-term exposures to
ozone
Health Category
Causal Determination
Respiratory Effects
Likely to be a causal relationship
Cardiovascular Effects
Suggestive of a causal relationship
Reproductive and Developmental Effects
Suggestive of a causal relationship
Central Nervous System Effects
Suggestive of a causal relationship
Carcinogenicity and Genotoxicity
Inadequate to infer a causal relationship
Mortality
Suggestive of a causal relationship
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8 POPULATIONS POTENTIALLY AT INCREASED
RISK FOR OZONE-RELATED HEALTH EFFECTS
1 Interindividual variation in human responses to air pollution exposure suggests that some
2 groups are at increased risk for detrimental effects in response to ambient exposure to an
3 air pollutant. The NAAQS are intended to provide an adequate margin of safety for both
4 the population as a whole and those individuals potentially at increased risk for health
5 effects in response to ambient air pollution (Preface to this ISA). To facilitate the
6 identification of populations at greater risk for O3-related health effects, studies have
7 evaluated factors that may contribute to the susceptibility and/or vulnerability of an
8 individual to O3. The definitions of susceptibility and vulnerability have been found to
9 vary across studies, but in most instances "susceptibility" refers to biological or intrinsic
10 factors (e.g., lifestage, sex) while "vulnerability" refers to nonbiological or extrinsic
11 factors (e.g., socioeconomic status [SES]) (U.S. EPA. 2010c. 2009d). Additionally, in
12 some cases, the terms "at-risk" and "sensitive" populations have been used to encompass
13 these concepts more generally. Previous IS As and reviews (Sacks etal.. 2011; U.S. EPA.
14 2010c. 2009d) have used an all encompassing definition for "susceptible population" to
15 focus on identifying the populations at greater risk for O3-induced heath effects and
16 circumvent the need to distinguish between susceptible and vulnerable factors. In this
17 chapter, "at-risk" groups are defined as those with characteristics that increase the risk of
18 O3-related health effects in a population. These characteristics include various factors,
19 such as genetic background, race, sex, lifestage, diet, preexisting disease, SES, and
20 characteristics that may modify exposure to O3 (e.g., time spent outdoors).
21 Individuals, and ultimately populations, could experience increased risk for O3-induced
22 health effects due to:
23 • Intrinsically increased risk: This describes individuals at greater risk due to a
24 biological mechanism;
25 • Extrinsically increased risk: This describes individuals at greater risk due to an
26 external, non-biological factor; and
27 • Increased dose: This describes individuals that have a greater dose at a given
28 concentration due to breathing patterns or other factors
29 In addition, some individuals might be placed at risk of experiencing a greater exposure
30 by being exposed at higher concentrations. For example, individuals in lower SES groups
31 might be exposed to higher O3 concentrations due to less availability/use of home air
32 conditioners (i.e., more open windows on high O3 days).
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1 Examples of potential factors intrinsically related to increased risk through biological
2 mechanisms are genetic background and sex, while extrinsic factors, such as SES, may
3 also increase the risk of O3-related health effects. However, some factors that may lead to
4 increased risk of O3-related health effects have both intrinsic and extrinsic components.
5 For example, SES may affect access to medical care, which could then affect the
6 presence of preexisting diseases and conditions often considered to be intrinsic factors.
7 Additionally, children tend to spend more time outdoors than adults, which increases
8 their dose of O3, but they also have intrinsic differences compared to adults based on
9 lung growth and development.
10 The emphasis of this chapter is on identifying and understanding the characteristics that
11 potentially increase the risk of O3-related health effects, regardless of whether the
12 increased risk at a given concentration is due to intrinsic factors, extrinsic factors, or
13 increased dose. The following sections examine factors that may modify the association
14 between O3 and health effects, but does not categorize them as intrinsic factors, extrinsic
15 factors, or increased dose, due to the convoluted and often connected pathways between
16 factors. However, the different role of intrinsic risk, extrinsic risk, and increased dose are
17 discussed as appropriate throughout the chapter.
18 Epidemiologic studies often conduct stratified analyses to identify the presence or
19 absence of effect measure modification to indicate whether O3 differentially affects
20 certain populations. This allows researchers to examine the effects of O3 exposure within
21 each group under study. A thorough evaluation of potential effect measure modifiers may
22 help identify populations that are more at-risk to health effects associated with O3
23 exposure. Toxicological and controlled human exposure studies can provide support and
24 biological plausibility for factors that may lead to increased risk for O3-related health
25 effects through the study of animal models of disease or individuals with underlying
26 disease or genetic polymorphisms that allow for comparisons between subgroups. The
27 results from these studies, combined with those results obtained through stratified
28 analyses in epidemiologic studies, comprise the overall weight of evidence for the
29 increased risk of specific populations to O3-related health effects.
30 This chapter discusses the epidemiologic, controlled human exposure, and toxicological
31 studies evaluated in Chapters 5, 6, and 7 that provide information on potential at-risk
32 populations. The epidemiologic studies included in this chapter consist of only those
33 studies that presented stratified results (e.g., males versus females or <65 years of age
34 versus > 65 years of age). This approach allowed for a comparison between populations
35 exposed to similar O3 concentrations and within the same study design. Numerous
36 studies that focus on only one potentially at-risk population are described in previous
37 chapters, but these studies are not discussed in detail in this chapter because of the lack of
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1 an adequate comparison group within the study. Controlled human exposure studies that
2 consisted of individuals with an underlying disease or genetic polymorphism, or studies
3 that categorized the study population by age, race, etc., and toxicological studies that
4 used animal models of disease were also evaluated for coherence and biological
5 plausibility.
6 Factors examined that may lead to increased risk of O3-related health effects based on the
7 overall evidence integrated across disciplines are described in greater detail in the
8 following sections.
8.1 Preexisting Disease/Conditions
9 Individuals with certain preexisting diseases are likely to constitute an at-risk population.
10 Previous O3 AQCDs concluded that some people with preexisting pulmonary disease,
11 especially asthma, are among those at increased risk from O3 exposure. Extensive
12 toxicological evidence was available indicating that altered physiological, morphological
13 and biochemical states typical of respiratory diseases, such as asthma, COPD, and
14 chronic bronchitis, may render people sensitive to additional oxidative burden induced by
15 O3 exposure. In addition, a number of epidemiologic studies found that some individuals
16 with respiratory diseases are at increased risk of O3-related effects. Little evidence,
17 however, was available on the potential for increased risk for people with other
18 preexisting conditions, such as cardiovascular diseases.
19 Recent studies that examined whether preexisting diseases and conditions lead to
20 increased risk of O3-induced health effects were identified and are summarized below.
21 Table 8-1 displays the prevalence rates of some of these conditions categorized by age
22 and region among adults in the U.S. population; data for children, when available, are
23 presented within sections. Substantial proportions of the U.S. population are affected by
24 these conditions and therefore may represent a potentially large at-risk population. While
25 these diseases and conditions are intrinsic to individuals, the pathways to their
26 development may have intrinsic or extrinsic origins.
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Table 8-1 Prevalence of respiratory diseases, cardiovascular diseases, and
diabetes among adults by age and region in the U.S.
Adults
Chronic Disease/Condition
N (in thousands)
Age
18-44
45-64
65-74
75+
Region
Northeast
Midwest
South
West
Respiratory Diseases
Asthma8
28,260
13.5
12.0
12.0
10.0
12.8
13.4
11.2
13.9
COPD
Chronic Bronchitis
Emphysema
9,832
3,789
3.2
0.2
5.5
2.0
5.9
5.7
5.3
5.0
3.4
1.2
4.8
1.9
5.2
1.9
2.9
1.3
Cardiovascular Diseases
All Heart Disease
Coronary Heart Disease
Hypertension
Diabetes
26,628
14,428
56,159
18,651
4.6
1.1
8.7
2.3
12.3
6.7
32.5
12.1
26.7
16.9
54.4
20.4
39.2
26.7
61.1
17.3
11.3
5.7
22.9
4.5
12.7
6.5
24.1
7.6
12.2
7.3
27.1
9.0
9.9
4.9
20.6
7.7
'Asthma prevalence is reported for "ever had asthma"
Source: Statistics for adults: Pleis et al. (2009)
8.1.1 Influenza/Infections
1 Recent studies have indicated that underlying infections may increase the risk of
2 individuals to O3-related health effects, although there are only a limited number of
3 studies. A study of hospitalizations in Hong Kong reported that increased levels of
4 influenza intensity resulted in increased excess risk of respiratory disease hospitalizations
5 related to O3 exposure (Wong et al.. 2009). In addition, a study of lung function in
6 asthmatic children reported decreases in lung function with increased short-term O3
7 exposure for those with upper respiratory infections but not for those without infections
8 (Lewis et al.. 2005). Toxicological studies provide biological plausibility for the increase
9 in Os-induced health effects observed in epidemiologic studies that examined infections.
10 Toxicological studies demonstrated that 0.08 ppm O3 increased streptococcus-induced
11 mortality, regardless of whether O3 exposure precedes or follows infection (Miller et al..
12 1978; Coffin and Gardner. 1972; Coffin et al., 1967). Ozone exposure likely impairs host
13 defenses, which may increase mortality due to an infectious agent. However, there is little
14 toxicological evidence that infection or influenza itself renders an individual at greater
15 risk of an O 3 -induced health effect.
8.1.2 Asthma
16 Previous O3 AQCDs identified individuals with asthma as a population at risk for O3-
17 related health effects. Within the U.S., approximately 12% of adults have reported ever
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1 having asthma (Pleis et al., 2009). The prevalence of asthma is approximately 7.2%.
2 16.2%, and 16.6% among U.S. children aged 0-4, 5-11, and 12-17, respectively (Bloom
3 et al.. 2008).
4 Multiple epidemiologic studies included within this ISA have evaluated the potential for
5 increased risk of O3-related health effects among individuals with asthma. A study of
6 lifeguards in Texas reported decreased lung function with short-term O3 exposure among
7 both individuals with and without asthma, however, the decrease was greater among
8 those with asthma (Thaller et al., 2008). A Mexican study of children ages 6-14 detected
9 an association between short-term O3 and wheeze, cough, and bronchodilator use among
10 asthmatics but not non-asthmatics, although this may have been the result of a small
11 non-asthmatic population (Escamilla-Nunez et al.. 2008). A study of the modification of
12 the effect of greater O3 associated decreases in short-term O3 exposure on lung function
13 by airway hyperresponsiveness (AHR) (a condition common among asthmatics) reported
14 greater O3-associated decreases in lung function in elderly individuals with AHR,
15 especially among those who were obese (Alexeeff et al.. 2007). However, no evidence
16 for increased risk was found in a study performed among children in Mexico City that
17 examined the effect of short-term O3 exposure on respiratory health (Barraza-Villarreal et
18 al.. 2008). In this study, a positive association was reported for airway inflammation
19 among asthmatic children, but the observed association was similar in magnitude to that
20 of non-asthmatics. Similarly, a study of children in California reported an association
21 between O3 concentration and exhaled nitric oxide fraction (FeNO) that persisted both
22 among children with and without asthma as well as those with and without respiratory
23 allergy (Berhane etal., 2011). Finally, some studies have reported null results for both
24 individuals with and without asthma. Khatri et al. (2009) found no association between
25 short-term O3 exposure and altered lung function for either asthmatic or non-asthmatic
26 adults, but did note a decrease in lung function among individuals with allergies.
27 Additional evidence for difference in effects among asthmatics has been observed in
28 studies that examined the association between O3 exposure and altered lung function by
29 asthma medication use. A study of children with asthma living in Detroit reported a
30 greater association between short-term O3 and lung function for corticosteroid users
31 compared with noncorticosteroid users (Lewis et al.. 2005). Conversely, another study
32 found decreased lung function among noncorticosteroid users compared to users,
33 although in this study, a large proportion of non-users were considered to be persistent
34 asthmatics (Hernandez-Cadena et al.. 2009). Lung function was not related to short-term
35 O3 exposure among corticosteroid users and non-users in a study taking place during the
36 winter months in Canada (Liu et al., 2009a). Additionally, a study of airway
37 inflammation reported a counterintuitive inverse association with O3 of similar
38 magnitude for all groups of corticosteroid users and non-users (Qian et al.. 2009).
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1 Controlled human exposure studies that have examined the effects of O3 on both
2 individuals with asthma and healthy controls are limited. Based on studies reviewed in
3 the 1996 and 2006 O3 AQCDs, subjects with asthma appeared to be more sensitive to
4 acute effects of O3 in terms of FEVi and inflammatory responses than healthy
5 non-asthmatic subjects. For instance, Horstman et al. (1995) observed that
6 mild-to-moderate asthmatics, on average, experienced double the O3-induced FEVi
7 decrement of healthy subjects (19% versus 10%, respectively, p = 0.04). Moreover, a
8 statistically significant positive correlation between FEVi responses to O3 exposure and
9 baseline lung function was observed in individuals with asthma, i.e., responses increased
10 with severity of disease. Minimal evidence exists suggesting that individuals with asthma
11 have smaller O3-induced FEVi decrements than healthy subjects (3% versus 8%,
12 respectively) (Mudway et al.. 2001). However, the asthmatics in that study also tended to
13 be older than the healthy subjects, which could partially explain their lesser response
14 since FEVi responses to O3 exposure diminish with age. Individuals with asthma also
15 had significantly more neutrophils in the BALF (18 hours postexposure) than similarly
16 exposed healthy individuals (Pedenetal.. 1997; Scannell et al.. 1996; Bashaetal.. 1994).
17 Furthermore, one newer study examined the effects of O3 on both individuals with atopic
18 asthma and healthy controls (Hernandez et al.. 2010). Greater numbers of neutrophils,
19 higher levels of cytokines and hyaluronan, and greater expression of macrophage
20 cell-surface markers were observed in induced sputum of atopic asthmatics compared
21 with healthy controls. Differences in O3-induced epithelial cytokine expression were
22 noted in bronchial biopsy samples from asthmatics and healthy controls (Bosson et al..
23 2003). Cell-surface marker and cytokine expression results, and the presence of
24 hyaluronan, are consistent with O3 having greater effects on innate and adaptive
25 immunity in these asthmatic individuals (see Section 5.4.2.2). In addition, older studies
26 have demonstrated that O3 exposure leads to increased bronchial reactivity to inhaled
27 allergens in mild allergic asthmatics (Kehrl et al.. 1999; Torres et al.. 1996) and to the
28 influx of eosinophils in individuals with pre-existing allergic disease (Vagaggini et al..
29 2002; Pedenetal.. 1995). Taken together, these results point to several mechanistic
30 pathways which could account for the enhanced sensitivity to O3 in subjects with asthma
31 (see Section 5.4.2.2).
32 Toxicological studies provide biological plausibility for greater effects of O3 among
33 those with asthma or AHR. In animal toxicological studies, an asthmatic phenotype is
34 modeled by allergic sensitization of the respiratory tract. Many of the studies that provide
35 evidence that O3 exposure is an inducer of AHR and remodeling utilize these types of
36 animal models. For example, a series of experiments in infant rhesus monkeys have
37 shown these effects, but only in monkeys sensitized to house dust mite allergen (Fanucchi
38 et al.. 2006: Joad et al.. 2006: Schelegle et al.. 2003). Similarly, Funabashi et al. (2004)
39 demonstrated adverse changes in pulmonary function in mice exposed to O3, and Wagner
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1 et al. (2007) demonstrated enhanced inflammatory responses in rats exposed to O3, but
2 only in animals sensitized to allergen. In general, it is the combined effects of O3 and
3 allergic sensitization which result in measurable effects on pulmonary function. In a
4 bleomycin induced pulmonary fibrosis model, exposure to 250 ppb O3 for 5 days
5 increased pulmonary inflammation and fibrosis, along with the frequency of
6 bronchopneumonia in rats. Thus, short-term exposure to O3 may enhance damage in a
7 previously injured lung (Oyarzun et al., 2005).
8 In the 2006 O3 AQCD, the potential for individuals with asthma to have greater risk of
9 O3-related health effects was supported by a number of controlled human exposure
10 studies, evidence from toxicological studies, and a limited number of epidemiologic
11 studies. Overall, in the recent epidemiologic literature some, but not all, studies report
12 greater risk of health effects among individuals with asthma. Studies examining effect
13 measure modification of the relationship between short-term O3 exposure and altered
14 lung function by corticosteroid use provided limited evidence of O3-related health
15 effects. Inconsistent findings observed in epidemiologic studies may be due to the
16 differences in O3 concentration across the studies. Additionally, recent studies of
17 behavioral responses have found that studies do not take into account individual
18 behavioral adaptations to forecasted air pollution levels (such as avoidance and reduced
19 time outdoors), which may underestimate the observed associations in studies that
20 examined the effect of O3 exposure on respiratory health (Neidell and Kinney. 2010).
21 This could explain some inconsistency observed among recent epidemiologic studies.
22 The evidence from controlled human exposure studies provides support for increased
23 detriments in FEVi and greater inflammatory responses to O3 in individuals with asthma
24 than in healthy individuals without a history of asthma. The collective evidence for
25 increased risk of O3-related health effects among individuals with asthma from controlled
26 human exposure studies is supported by recent toxicological studies which provide
27 biological plausibility for heightened risk of asthmatics to respiratory effects due to O3
28 exposure.
8.1.3 Chronic Obstructive Pulmonary Disease (COPD)
29 Although not extensively examined in the literature, initial evidence suggests that
30 preexisting COPD may modify the association between short-term O3 exposure and
31 cardiovascular-related health effects. In the U.S. over 4% of adults report having chronic
32 bronchitis and almost 2% report having emphysema, both of which are classified as
33 COPD (Pleis et al.. 2009).
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1 In a recent study, Peel et al. (2007) found that individuals with COPD were at increased
2 risk of cardiovascular ED visits in response to short-term O3 exposure compared to
3 healthy individuals in Atlanta, GA. The authors reported that short-term O3 exposure was
4 associated with higher odds of an ED visit for peripheral and cerebrovascular disease
5 among individuals with COPD compared to individuals without COPD. However,
6 preexisting COPD did not increase the odds of hospitalization for all CVD outcomes (i.e.
7 IHD, dysrhythmia, or congestive heart failure). In an additional study performed in
8 Taiwan, both individuals with and without COPD had higher odds of congestive heart
9 failure associated with O3 exposure on warm days (Lee et al.. 2008a). An additional
10 study also found no association between O3 exposure and lung function regardless of
11 whether the study participant had COPD or other health issues (asthma or IHD) (Lagorio
12 et al.. 2006).
13 Recent epidemiologic evidence indicates that persons with COPD may have increased
14 O3-related cardiovascular effects, but little information is available for other O3-related
15 health effects among individuals with COPD.
8.1.4 Cardiovascular Disease
16 Cardiovascular disease (CVD) has become increasingly prevalent in the U.S., with about
17 12% of adults reporting a diagnosis of heart disease (Table 8-1). A high prevalence of
18 other cardiovascular-related conditions has also been observed, such as hypertension
19 which is prevalent among approximately 24% of adults. In the 2006 O3 AQCD, little
20 evidence was available regarding preexisting CVD as a susceptibility factor. Recent
21 epidemiologic studies have examined cardiovascular-related diseases as modifiers of the
22 O3-outcome associations; however, no recent evidence is available from controlled
23 human exposure studies or toxicological studies.
24 Peel et al. (2007) compared the associations between short-term O3 exposure and
25 cardiovascular ED visits in Atlanta, GA among multiple comorbid conditions. The
26 authors found no evidence of increased risk of cardiovascular ED visits in individuals
27 previously diagnosed with dysrhythmia, congestive heart failure, or hypertension
28 compared to healthy individuals. Similarly, a study in France examined the association
29 between O3 concentrations and ischemic cerebrovascular events (ICVE) and myocardial
30 infarction (MI) and the influence of multiple vascular risk factors on any observed
31 associations (Henrotin et al.. 2010). The association between O3 exposure and ICVE was
32 elevated for individuals with multiple risk factors, specifically individuals with diabetes
33 or hypertension. For the association between O3 and MI, increased odds were apparent
34 only for those with hypercholesterolaemia. In a study conducted in Taiwan, a positive
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1 association was observed for O3 on warm days and congestive heart failure hospital
2 admissions (HAs), but the association did not differ between individuals with/without
3 hypertension or with/without dysrhythmia (Lee et al.. 2008a). Another study in Taiwan
4 reported that the association between O3 levels and ED visits for arrhythmias were
5 greater on warm days among those with congestive heart failure compared to those
6 without congestive heart failure; however, the estimate and 95% CIs for those without
7 congestive heart failure is completely contained within the 95% CI of those with
8 congestive heart failure (Chiu and Yang. 2009).
9 Although not studied extensively, a study has examined the increased risk of O3-related
10 changes in blood markers for individuals with CVD. There was a greater association
11 between O3 exposure and some, but not all, blood inflammatory markers among
12 individuals with a history of CVD. Liao et al. (2005) found that fibrinogen was positively
13 associated with short-term O3 exposure but this association was present only among
14 individuals with a history of CVD. No association was observed among those without a
15 history of CVD. However, for another biomarker (vWF), CVD status did not modify the
16 positive association with short-term O3 exposure (Liao et al.. 2005).
17 Mortality studies provide some evidence for a potential increase in O3-induced mortality
18 in individuals with preexisting atrial fibrillation and atherosclerosis. In a study of 48 U.S.
19 cities, increased risk of mortality with short-term O3 exposure was observed only among
20 individuals with secondary atrial fibrillation (Medina-Ramon and Schwartz. 2008). No
21 association was observed for short-term O3 exposure and mortality in a study of
22 individuals with diabetes with or without CVD prior to death; however, there was some
23 evidence of increased risk of mortality during the warm season if individuals had diabetes
24 and atherosclerosis compared to only having diabetes (Goldberg et al.. 2006).
25 Finally, although not extensively examined, a study explored whether a preexisting CVD
26 increased the risk of an O3-induced respiratory effect. Lagorio et al. (2006) examined the
27 effect of O3 exposure on lung function among participants with a variety of preexisting
28 diseases, including IHD. No association was observed regardless of whether the
29 participant had IHD.
30 Overall, most short-term exposure studies did not report increased O3-related health
31 effects for individuals with preexisting CVD, with the possible exception of O3 exposure
32 and mortality. Future research among those with CVD compared to those without will
33 increase the understanding of potential increased risk of O3-related health effects among
34 this group.
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8.1.5 Diabetes
1 Recent literature has not extensively examined whether individuals with diabetes (about
2 8% of U.S. adults) are potentially at increased risk of O3-related health effects. In a study
3 of short-term O3 exposure and cardiovascular ED visits in Atlanta, GA, no association
4 was observed for individuals with or without diabetes (Peel et al., 2007). A similar study
5 conducted in Taiwan reported a positive association between O3 exposure on warm days
6 and HAs for congestive heart failure; however, no modification of the association by
7 diabetes was observed (Lee et al.. 2008a). Finally, in a study of O3 exposure and ED
8 visits for arrhythmia in Taiwan, there was no evidence of effect measure modification by
9 diabetes on warm or cool days (Chiu and Yang. 2009).
8.1.6 Hyperthyroidism
10 Hyperthyroidism has been identified in toxicological studies as a potential factor that may
11 lead to increased risk of O3-related health effects but has not yet been explored in
12 epidemiologic or controlled human exposure studies. Lung damage and inflammation due
13 to oxidative stress may be modulated by thyroid hormones. Compared to controls,
14 hyperthyroid rats exhibited elevated levels of BAL neutrophils and albumin after a 4-hour
15 exposure to O3, indicating O3-induced inflammation and damage. Hyperthyroidism did
16 not affect production of reactive oxygen or nitrogen species, but BAL phospholipids were
17 increased, indicating greater activation of Type II cells and surfactant protein production
18 compared to normal rats (Huffman et al.. 2006). Thus, this study provides some
19 underlying evidence which suggests that individuals with hyperthyroidism may represent
20 an at-risk population.
8.2 Lifestage
21 The 1996 and 2006 O3 AQCDs identified children, especially those with asthma, and
22 older adults as at-risk populations. These previous AQCDs reported clinical evidence that
23 children have greater spirometric responses to O3 than middle-aged and older adults
24 (U.S. EPA. 1996a). Similar results were observed for symptomatic responses and O3
25 exposure. Among older adults, most studies reported in the 2006 O3 AQCD reported
26 greater effects of short-term O3 exposure and mortality compared to other age groups.
27 New evidence, summarized below, further supports these findings.
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8.2.1 Children
1 The 2000 Census reported that 28.6% of the U.S. population was under 20 years of age,
2 with 14.1% under the age of 10 (SSDAN CensusScope. 2010a). Children are considered
3 to be more at risk for O3-related health effects compared to adults because they spend
4 more time outside and are more highly active, especially during the summer when O3
5 concentrations are the highest (U.S. EPA. 2006b). Moreover, children's respiratory
6 systems are undergoing lung growth until about 18-20 years of age and are therefore
7 thought to be intrinsically more at risk for O3-induced damage (U.S. EPA. 2006b).
8 The 1996 O3 AQCD, reported clinical evidence that children, adolescents, and young
9 adults (<18 years of age) appear, on average, to have nearly equivalent spirometric
10 responses to O3 exposure, but have greater responses than middle-aged and older adults
11 (U.S. EPA. 1996a). Sycalmptomatic responses (e.g., cough, shortness of breath, pain on
12 deep inspiration) to O3 exposure, however, appear to increase with age until early
13 adulthood and then gradually decrease with increasing age (U.S. EPA. 1996a). For
14 subjects aged 18-36 years, McDonnell et al. (1999) reported that symptom responses
15 from O3 exposure also decrease with increasing age. Complete lung growth and
16 development is not achieved until 18-20 years of age in women and the early 20s for
17 men; pulmonary function is at its maximum during this time as well. Additionally, PBPK
18 modeling reported regional extraction of O3 to be higher in infants compared to adults.
19 This is thought to be due to the smaller nasal and pulmonary regions' surface area in
20 children under the age of 5 years compared to the total airway surface area observed in
21 adults (Sarangapani et al., 2003).
22 Recent epidemiologic studies have been performed examining different age groups and
23 their susceptibility to O3-related respiratory HAs and emergency department (ED) visits.
24 A study in Cyprus of short-term O3 concentrations and respiratory HA detected possible
25 effect measure modification by age with a larger association among individuals <
26 15 years of age compared with those > 15 years of age. However, this difference was
27 only apparent with a 2-day lag (Middleton et al., 2008). Similarly, a Canadian study of
28 asthma-ED visits reported the strongest O3-related associations among 5- to 14-year olds
29 compared to the other age groups (ages examined 0-75+) (Villeneuve et al., 2007).
30 Greater O3-associated change in asthma-related ED visits were also reported among
31 children (<15 years) as compared to adults (15 to 64 years) in a study from Finland
32 (Halonen et al.. 2009). A study of New York City HAs demonstrated an increase in the
33 association between O3 exposure and asthma-related HAs for 6- to 18-year olds
34 compared to those < 6 years old and those > 18 years old (Silverman and Ito. 2010).
35 When examining long-term O3 exposure and asthma HA among children, associations
36 were determined to be larger among children 1 to 2 years old compared to children 2 to 6
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1 years old (Lin et al., 2008b). A few studies reported positive associations among both
2 children and adults and no modification of the effect by age. A study performed in Hong
3 Kong examined O3 exposure and asthma-related HAs for ages 0 to!4, 15 to 65, and >65
4 (Ko et al.. 2007). The researchers reported that the association was greater among the 0 to
5 14 and 14 to 65 age groups compared to the >65 age group. Another study looking at
6 asthma-related ED visits and O3 exposure in Maine reported positive associations for all
7 age groups (ages 2 to 65) (Paulu and Smith. 2008). Effects of O3 exposure on asthma
8 hospitalizations among both children and adults (<18 and >18 years old) were
9 demonstrated in a study in Washington, but only children (<18 years of age) had
10 statistically significant results at lag day 0, which the authors wrote, "suggests that
11 children are more immediately responsive to adverse effects of O3 exposure" (Mar and
12 Koenig. 2009).
13 The evidence observed in epidemiologic studies is supported by recent toxicological
14 studies which observed O3-induced health effects in immature animals. Early life
15 exposures of multiple species of laboratory animals, including infant monkeys, resulted
16 in changes in conducting airways at the cellular, functional, ultra-structural, and
17 morphological levels. Carey et al. (2007) conducted a study of O3 exposure in infant
18 rhesus macaques, whose nasal airways closely resemble that of humans. Monkeys were
19 exposed either acutely for 5 days to 0.5 ppm O3, or episodically for 5 biweekly cycles
20 alternating 5 days of 0.5 ppm O3 with 9 days of filtered air, designed to mimic human
21 exposure (70 days total). All monkeys acutely exposed to O3 had moderate to marked
22 necrotizing rhinitis, with focal regions of epithelial exfoliation, numerous infiltrating
23 neutrophils, and some eosinophils. The distribution, character, and severity of lesions in
24 episodically exposed monkeys were similar to that of acutely exposed animals. Neither
25 group exhibited mucous cell metaplasia proximal to the lesions, a protective adaptation
26 observed in adult monkeys exposed continuously to 0.3 ppm O3 in another study
27 (Harkema et al., 1987a). Functional (increased airway resistance and responsiveness with
28 antigen + O3 co-exposure) and cellular changes in conducting airways (increased
29 numbers of inflammatory eosinophils) also manifested among the infant monkeys
30 (Plopper et al.. 2007). In addition, the lung structure of the conducting airways was
31 significantly stunted or altered versus control animals and this aberrant development was
32 persistent 6 months postexposure (Fanucchi et al.. 2006).
33 Similarly, rat fetuses exposed to O3 in utero had significant ultrastructural changes in
34 bronchiolar epithelium when examined near the end of gestation (Lopez et al., 2008). In
35 addition, exposure of mice to mixtures of air pollutants early in development affected pup
36 lung cytokine levels (TNF, IL-1, KC, IL-6, and MCP-1) (Auten et al.. 2009). In utero
37 exposure of animals to PM augmented O3-induced airway hyper-reactivity in these pups
38 as juveniles.
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1 Age may affect the inflammatory response to O3 exposure. In comparing neonatal mice
2 to adult mice, increased bronchoalveolar lavage (BAL) neutrophils were observed in four
3 strains of neonates 24 hours after exposure to 0.8 ppm O3 for 5 hours (Vancza et al.,
4 2009). Three of these strains also exhibited increased BAL protein, although the two
5 endpoints were not necessarily consistently correlated in a given strain. In some strains,
6 however, adults were more sensitive, indicating a strain-age interaction. Toxicological
7 studies reported that the difference in effects among younger lifestage may be due to
8 age-related changes in endogenous antioxidants and sensitivity to oxidative stress. A
9 recent study demonstrated that 0.25 ppm O3 exposure differentially alters expression of
10 metalloproteinases in the skin of young (8 weeks old) and aged (18 months old) mice,
11 indicating age-related susceptibility to oxidative stress (Fortino et al., 2007). Valacchi et
12 al. (2007) found that aged mice had more vitamin E in their plasma but less in their lungs
13 compared to young mice, which may affect their pulmonary antioxidant defenses. Servais
14 et al. (2005) found higher levels of oxidative damage indicators in immature (3 weeks
15 old) and aged (20 months old) rats compared to adult rats, which were relatively resistant
16 to an intermittent 7-day exposure to 0.5 ppm O3. Immature rats exhibited a higher
17 ventilation rate, which may have increased exposure. Additionally, a series of
18 toxicological studies reported an association between O3 exposure and bradycardia that
19 was present among young mice but not among older mice (Hamade et al.. 2010;
20 Tankersley et al.. 2010; Hamade and Tankersley. 2009; Hamade et al., 2008). Regression
21 analysis revealed a significant interaction between age and strain on heart rate, which
22 implies that aging may affect heart rate differently between mouse strains (Tankersley et
23 al.. 2010). The authors proposed that the genetic differences between the mice strains
24 could be altering the formation of ROS, which tends to increase with age, thus
25 modulating the changes in cardiopulmonary physiology after O3 exposure.
26 The previous and current human clinical and toxicological studies reported evidence of
27 increased risk from O3 exposure for younger ages, which provides coherence and
28 biological plausibility to the epidemiologic studies on children. Recent studies of
29 respiratory HAs and ED visits observed inconsistent findings for associations among
30 children and young adults, although generally studies reported positive associations
31 among both children and adults or just among children. For other outcomes, there were
32 also inconsistent findings regarding increased risk of O3-related health effects. The
33 interpretation of these studies is limited by the lack of consistency in comparison age
34 groups and outcomes examined.
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8.2.2 Older Adults
1 Older adults may be at greater risk of health effects associated with O3 exposure through
2 a variety of intrinsic pathways. The gradual decline in physiological processes that occur
3 with aging may lead to increased risk of O3-related health effects (U.S. EPA. 2006a).
4 Older adults may also differ in amounts of exposure because diminished symptomatic
5 responses may allow the elderly to withstand increased continued O3 exposure. In
6 addition, older adults, in general, have a higher prevalence of preexisting diseases
7 compared to younger age groups and this may also lead to increased susceptibility to
8 O3-related health effects (see Table 8-1 that gives preexisting rates by age).With the
9 number of older Americans increasing in upcoming years (estimated to increase from
10 12.4% of the U.S. population to 19.7% between 2000 to 2030, which is approximately 35
11 million and 71.5 million individuals, respectively) this group represents a large
12 population potentially at risk of O3 -related health effects (SSDAN CensusScope, 2010a;
13 U.S. Census Bureau. 2010).
14 Multiple epidemiologic studies of O3 exposure and HAs were stratified by age groups. A
15 positive association was reported between O3 levels and respiratory HAs for adults >65
16 years old but not for those adults aged 15 to 64 years (Halonen et al.. 2009). In the same
17 study, no association was observed between O3 concentration and respiratory mortality
18 among those >65 years old or those 15 to 64 years old; however, an inverse association
19 between O3 concentration and cardiovascular mortality was present among individuals >
20 65 years old but not among individuals < 65 years old. This inverse association among
21 those >65 years old persisted when examining HAs for coronary heart disease. A study of
22 CVD-related hospital visits in Bangkok, Thailand reported an increase in percent change
23 for hospital visits with previous day and cumulative 2-day O3 levels among those >
24 65 years old, whereas no association was present for individuals less than 65 years of age
25 (Buadong et al.. 2009). No association was observed for current day or cumulative 3-day
26 averages in any age group. A study examining O3 and HAs for CVD-related health
27 effects reported no association for individuals aged 15 to 64 or individuals aged > 65
28 years, although one lag-time did show an inverse effect for coronary heart disease among
29 elderly that was not present among 15- to 64-year olds (Halonen et al.. 2009). No
30 modification by age (40 to 64 year olds versus >64 year olds) was observed in a study
31 from Brazil examining O3 levels and COPD ED visits (Arbex et al.. 2009).
32 The majority of recent studies reported greater effects of short-term O3 exposure and
33 mortality among older adults, which is consistent with the findings of the 2006 O3
34 AQCD. A study conducted in 48 cities across the U.S. reported larger effects among
35 adults >65 years old compared to those < 65 years (Medina-Ramon and Schwartz. 2008).
36 Further investigation of this study population revealed no association between O3
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1 exposure and mortality until age 50 and a reduced effect after age 80 (Zanobetti and
2 Schwartz. 2008a). A study of 7 urban centers in Chile reported similar results, with
3 greater effects in adults >65 years old, however the effects were smaller among those
4 >85 years old compared to those in the 75 to 84 years old age range (Cakmak et al..
5 2007). More recently, a study conducted in the same area reported similar associations
6 between O3 exposure and mortality in adults aged < 64 years old and 65 to 74 years old,
7 but the risk was increased among older age groups (Cakmak et al., 2011). A study
8 performed in China reported greater effects in populations >45 years old (compared to 5
9 to 44 year olds), with statistically significant effects present only among those >65 years
10 old (Kanetal.. 2008). An Italian study reported higher risk of all-cause mortality
11 associated with increased O3 concentrations among individuals >85 year old as compared
12 to those 35 to 84 years old. Those 65 to 74 and 75 to 84 years old did not show a greater
13 increase in risk compared to those aged 35 to 64 years (Stafoggia et al.. 2010). The Air
14 Pollution and Health: A European and North American Approach (APHENA) project
15 examined the association between O3 exposure and mortality for those <75 and >
16 75 years of age. In Canada, the associations for all-cause and cardiovascular mortality
17 were greater among those >75 years old in the summer-only and all-year analyses. Age
18 groups were not compared in the analysis for respiratory mortality in Canada. In the U.S.,
19 the association for all-cause mortality was slightly greater for those <75 years of age
20 compared to those >75 years old in summer-only analyses. No consistent pattern was
21 observed for CVD mortality. In Europe, slightly larger associations for all-cause
22 mortality were observed in those <75 years old in all-year and summer-only analyses.
23 Larger associations were reported among those <75years for CVD mortality in all-year
24 analyses, but the reverse was true for summer-only analyses (Katsouyanni et al.. 2009).
25 Biological plausibility for increased risk among older adults is provided by clinical and
26 toxicological studies. Respiratory symptom responses to O3 exposure appears to increase
27 with age until early adulthood and then gradually decrease with increasing age (U.S.
28 EPA. 1996a). which may put them at increased risk by withstanding continued O3
29 exposure. Regarding cardiac outcomes, biological plausibility is provided by a
30 toxicological study. O3 exposure resulted in an increase in left ventricular chamber
31 dimensions at end diastole (LVEDD) in young and old mice, whereas decreases in left
32 ventricular posterior wall thickness at end systole (PWTES) were only observed among
33 older mice (Tankersley et al., 2010). Other toxicological studies also indicate increased
34 susceptibility in older animals for some endpoints. The hippocampus, one of the main
35 regions affected by age-related neurodegenerative diseases, may be more sensitive to
36 oxidative damage in aged rats. In a study of young (47 days) and aged (900 days) rats
37 exposed to 1 ppm O3 for 4 hours, O3-induced lipid peroxidation occurred to a greater
38 extent in the striatum of young rats, whereas it was highest in the hippocampus in aged
39 rats (Rivas-Arancibia et al.. 2000). In young mice, healing of skin wounds is not
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1 significantly affected by O3 exposure (Lim et al., 2006). However, exposure to 0.5 ppm
2 O3 for 6 h/day significantly delays wound closure in aged mice.
3 Although some outcomes reported mixed findings regarding an increase in risk for older
4 adults, recent studies of O3 exposure and mortality reported associations present for older
5 adults. This is consistent with the results reported in the 2006 O3 AQCD.
8.3 Sex
6 The distribution of males and females in the U.S. is similar. In 2000, 49.1% of the U.S.
7 population was male and 50.9% were female. The distribution did vary by age with a
8 greater prevalence of females >65 years old compared to males (SSDAN CensusScope.
9 2010a). The 2006 O3 AQCD did not report evidence of differences between the sexes in
10 health responses to O3 exposure. Recent epidemiologic studies have evaluated the effects
11 of short-term and long-term exposure to O3 on multiple health endpoints stratified by sex
12 and overall, the results are inconsistent.
13 A study in Maine that examined short-term O3 concentrations and asthma ED visits
14 detected greater effects among males ages 2 to!4 years and among females ages 15 to 34
15 years compared to males and females in the same age groups (no difference was detected
16 for males and females aged 35 to 64) (Paulu and Smith. 2008). A Canadian study
17 reported no associations between short-term O3 and respiratory infection HAs for either
18 boys or girls under the age of 15 (Lin et al., 2005). whereas another Canadian study
19 reported a slightly higher but non-statistically significant increase in respiratory HA for
20 males (mean ages 47.6 to 69.0 years) (Cakmak et al., 2006b). A recent study from Hong
21 Kong examining individuals of all ages reported no effect measure modification by sex
22 for overall respiratory disease HAs, but did detect a greater excess risk of HAs for COPD
23 among females compared to males (Wong et al.. 2009). Similarly a study in Brazil found
24 higher effect estimates for COPD ED visits among females compared to males (Arbex et
25 al.. 2009). Higher levels of respiratory HA with greater O3 concentrations was also
26 observed for females in a study of individuals living in Cyprus (Middleton et al., 2008).
27 A study of lung function unrelated to HA and ED visits was conducted among lifeguards
28 in Texas and reported decreased lung function with increased O3 exposure among
29 females but not males (Thaller et al.. 2008). This study included individuals aged 16 to 27
30 years, and the majority of participants were male. A New York study found no effect
31 measure modification of the association between long-term O3 exposure and asthma HA
32 among males and females between 1 and 6 years old (Lin et al., 2008b).
33 In addition to examining the potential modification of O3 associations with respiratory
34 outcomes by sex, studies also examined cardiovascular-related outcomes specifically
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1 HAs and ED visits. All of these studies reported no effect modification by sex with some
2 studies reporting null associations for both males and females (Wong et al.. 2009;
3 Middleton et al., 2008; Villeneuve et al.. 2006a) and one study reporting a positive
4 associations for both sexes (Cakmak et al.. 2006a). A French study examining the
5 associations between O3 concentrations and risk of ischemic strokes (not limited to ED
6 visits or HAs) reported no association for either males or females with lags of 0, 2, or
7 3 days (Henrotin et al., 2007). A positive association was reported for males with a lag of
8 1 day, but this association was null for females. The authors noted that men in the study
9 had much higher rates of current and former smoking than women (67.4% versus 9.3%).
10 A biomarker study investigating the effects of O3 concentrations on high-sensitivity
11 C-reactive protein (hs-CRP), fibrinogen, and white blood cell (WBC) count, reported
12 observations for various lag times ranging from 0 to 7 days (Steinvil et al.. 2008). Most
13 of the associations were null for males and females although one association between O3
14 and fibrinogen was positive for males and null for females (lag day 4); however, this
15 positive association was null or negative when other pollutants were included in the
16 model. Only one study examining correlations between O3 levels and oxidative DNA
17 damage examined results stratified by sex. In this study Palli et al. (2009) reported
18 stronger correlations for males than females, both during short-term exposure (less than
19 30 days) and long-term exposure (0-90 days). However, the authors commented that this
20 difference could have been partially explained by different distributions of exposure to
21 traffic pollution at work.
22 A few studies have examined the association between short-term O3 concentrations and
23 mortality stratified by sex and in contrast with studies of other endpoints, were more
24 consistent in reporting elevated risks among females. These studies, conducted in the
25 U.S. (Medina-Ramon and Schwartz. 2008). Italy (Stafoggia et al.. 2010). and Asia (Kan
26 et al.. 2008). reported larger effect estimates in females compared to males. In the U.S.
27 study, the elevated risk of mortality among females was greater specifically among those
28 >60 years old (Medina-Ramon and Schwartz. 2008). However, a recent study in Chile
29 reported similar associations between O3 exposure and mortality among both men and
30 women (Cakmak et al.. 2011). One long-term O3 exposure study of respiratory mortality
31 stratified their results by sex and reported relative risks of 1.01 (95 % CI: 0.99, 1.04) for
32 males and 1.04 (95% CIs 1.03, 1.07) for females (Jerrett et al.. 2009).
33 Experimental research provided a further understanding of the underlying mechanisms
34 that may explain a possible differential risk in O3-related health effects among males and
35 females. Several studies have suggested that physiological differences between sexes
36 may predispose females to a greater susceptibility to O3. In females, lower plasma and
37 nasal lavage fluid (NLF) levels of uric acid (most prevalent antioxidant), the initial
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1 defense mechanism of O3 neutralization, may be a contributing factor (Housley et al.,
2 1996). Consequently, reduced absorption of O3 in the upper airways of females may
3 promote its deeper penetration. Dosimetric measurements have shown that the absorption
4 distribution of O3 is independent of sex when absorption is normalized to anatomical
5 dead space (Bush et al., 1996). Thus, a differential removal of O3 by uric acid seems to
6 be minimal. In general, the physiologic response of young healthy females to O3
7 exposure appears comparable to the response of young males (Hazucha et al., 2003). A
8 few studies have examined changes in O3 responses during various menstrual cycle
9 phases. Lung function response to O3 was enhanced during the follicular phase of the
10 menstrual cycle compared to the luteal phase in a small study of women (Fox et al..
11 1993). However, Seal et al. (1996) later reported no effect of menstrual cycle phase in
12 their analysis of responses from 150 women, but conceded that the methods used by Fox
13 et al. (1993) more precisely defined the menstrual cycle phase. Another study also
14 reported no difference in responses among females during the follicular and luteal phases
15 of their cycle (Weinmann et al.. 1995a). Additionally, in this study the responses in
16 women were comparable to those reported for men in the study. In a toxicological study,
17 small differences in effects by sex were seen in adult mice with respect to pulmonary
18 inflammation and injury after a 5-h exposure to 0.8 ppm O3, and although adult females
19 were generally more susceptible, these differences were strain-dependent, with some
20 strains exhibiting greater susceptibility in males (Vancza et al., 2009). The most obvious
21 sex difference was apparent in lactating females, which incurred the greatest lung injury
22 or inflammation among several of the strains.
23 Overall, results have varied, with recent evidence for increased risk for O3-related health
24 effects present for females in some studies and males in other studies. Most studies
25 examining the associations O3 and mortality report females to be at greater risk than
26 males. Little evidence is available regarding a difference between the sexes for other
27 outcomes. Inconsistent findings were reported on whether effect measure modification
28 exists by sex for respiratory and cardiovascular HAs and ED visits.
8.4 Genetics
29 Multiple studies that examined the effect of short- and long-term O3 exposure on
30 respiratory function have focused on whether various gene profiles modify the effect of
31 O3 on various health effects. A study of wheeze in infants reported larger associations
32 between short-term O3 exposure and wheeze and difficulty breathing in infants whose
33 mothers have asthma compared to infants of mothers without asthma, illustrating the
34 potential for genetics to play a role in O3-related health effects (Triche et al.. 2006).
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1 Multiple genes, including glutathione S-transferase Mu 1 (GSTM1) and tumor necrosis
2 factor-a (TNF-a) were evaluated in the 2006 O3 AQCD and found to have a "potential
3 role... in the innate susceptibility to O3." Studies performed since the 2006 O3 AQCD
4 have continued to examine the roles of GSTM1 and TNF-a on O3-related health effects
5 and have also examined other gene variants that may increase the risk of O3-related
6 health effects. Due to small sample sizes, many controlled human exposure studies are
7 limited in their ability to test genes with low frequency and therefore, some genes
8 important for O3-related health effects may not have been examined.
9 Epidemiologic studies that examined the effects of short-term exposure to O3 on lung
10 function included analyses of potential gene-environment interactions. Romieu et al.
11 (2006) reported an association between O3 and respiratory symptoms that were larger
12 among children with GSTM1 null or glutathione S-transferase P 1 (GSTP1) Val/Val
13 genotypes. However, results suggested that O3-associated decreases in lung function may
14 be greater among children with GSTP1 lie/lie or Ile/Val compared to GSTP1 Val/Val.
15 Alexeef et al. (2008) reported greater decreases in lung function among GSTP1 Val/Val
16 adults than those with other genotypes. In addition, they detected greater decreases in
17 lung function for adults with long GT dinucleotide repeats in heme-oxygenase-1
18 (HMOX1) promoters.
19 Several controlled human exposure studies have reported that genetic polymorphism of
20 antioxidant enzymes may modulate pulmonary function and inflammatory response to O3
21 challenge. It appears that healthy carriers of NAD(P)H quinone oxidoreductase 1 (NQO1)
22 wild type (wt) in combination with GSTM1 null genotype had greater decreases in lung
23 function parameters with exposure to O3 (Bergamaschi et al.. 2001). Adults with GSTM1
24 null only genotype did not show the same response to O3. In contrast, asthmatic children
25 with GSTM1 null genotype (Romieu et al.. 2004a) were reported to have greater
26 decreases in lung function in relation to O3 exposure. In a similar study, Vagaggini et al.
27 (2010) exposed mild-to-moderate asthmatics to O3 during moderate exercise. In subjects
28 with NQO1 wt and GSTM1 null, there was no evidence of changes in lung function or
29 inflammatory responses to O3. Kim et al. (2011) also recently conducted a study among
30 young adults, about half of whom were GSTMl-null and half of whom were
31 GSTM1-sufficient. They detected no difference in the FEVi responses to O3 exposure by
32 GSTM1 genotype.
33 In a study of healthy volunteers with GSTM1 sufficient (n=19; 24 ± 3) and GSTM1 null
34 (n=16; 25 ± 5) genotypes exposed to 400 ppb O3 for 2 hours with exercise, Alexis et al.
35 (2009) found genotype effects on inflammatory responses but not lung function responses
36 to O3. At 4 hours post O3 exposure, individuals with either GSTM1 genotype had
37 significant increases in sputum neutrophils with a tendency for a greater increase in
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1 GSTM1 sufficient than GSTM1 nulls. At 24 hours postexposure, neutrophils had
2 returned to baseline levels in the GSTM1 sufficient individuals. In the GSTM1 null
3 subjects, neutrophil levels increased from 4 to 24 hours and were significantly greater
4 than both baseline levels and levels at 24 hours in the GSTM1 sufficient individuals. In
5 addition, O3 exposure increased the expression of the surface marker CD 14 in airway
6 neutrophils of GSTM1 null subjects compared with GSTM1 sufficient subjects. CD 14
7 and TLR4 are co-receptors for endotoxin, and signaling through this innate immune
8 pathway has been shown to be important for a number of biological responses to O3
9 exposure in toxicological studies (Garantziotis et al., 2010; Hollingsworth et al., 2010;
10 Hollingsworth et al.. 2004; Kleeberger et al.. 2000). Alexis et al. (2009) also
11 demonstrated decreased numbers of airway macrophages at 4 and 24 hours following O3
12 exposure in GSTM1 sufficient subjects. Airway macrophages in GSTM1 null subjects
13 were greater in number and found to have greater oxidative burst and phagocytic
14 capability than those of GSTM1 sufficient subjects. Airway macrophages and dendritic
15 cells from GSTM1 null subjects exposed to O3 expressed higher levels of the surface
16 marker HLA-DR, again suggesting activation of the innate immune system. Since there
17 was no FA control in the Alexis et al. (2009) study, effects of the exposure other than O3
18 cannot be ruled out. In general, the findings between these studies are inconsistent and
19 additional, better-controlled studies are needed to clarify the influence of genetic
20 polymorphisms on O3 responsiveness in humans.
21 Several epidemiologic studies of long-term O3 exposure examined interactions with
22 different gene variants, including GSTP1, HMOX1, and TNF-a using data from the
23 Children's Health Study. A study among children reported a three-way interaction effect
24 between He 105 homozygotes of GSTP1, O3 exposure, and playing more than two team
25 sports, and new onset of asthma (Islam et al.. 2009). Additionally, Islam et al. found that
26 non-Hispanic white children with less than 23 repeats in the HMOX1 gene had decreased
27 risk of new-onset asthma (Islam et al.. 2008). ARG1 and ARG2 (encoded by arginases)
28 modification were examined for the association between genotypes and new-onset
29 asthma (Salam et al., 2009). Reduced asthma risk was observed among atopic children
30 living in high O3 concentration areas and having the ARG1 haplotypes. There was no
31 difference in risk for children with ARG2 haplotypes. A decreased risk of bronchitic
32 symptoms was observed among asthmatic children in low O3 concentration areas with
33 TNF-a variant G-308A (TNF-308GG genotype), a variant that may alter gene expression.
34 There was no decrease in risk for children with this TNF-a variant and living in areas
35 with high O3 concentrations. Additionally, this modification for high and low levels of
36 O3 was not present among non-asthmatic children (Lee et al.. 2009b). Wenten et al.
37 (2009) observed increased risk of respiratory-related school absences among children
38 with variants of catalase (CAT) and myeloperoxidase (MPO) genes, especially when the
39 children were living in high O3 concentration areas.
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1 In general, toxicological studies have reported differences in cardiac and respiratory
2 effects after O3 exposure among different mouse strains, which alludes to differential risk
3 among individuals due to genetic variability (Tankersley et al.. 2010; Chuang et al.. 2009;
4 Hamade and Tankersley. 2009; Hamade et al.. 2008). Thus strains of mice which are
5 prone to or resistant to O3-induced effects have been used to systematically identify
6 candidate genes that may increase risk of O3-related health effects. Genome wide linkage
7 analyses have identified quantitative trait loci for O3-induced lung inflammation and
8 hyperpermeability on chromosome 17 (Kleeberger et al.. 1997) and chromosome 4
9 (Kleeberger et al., 2000). respectively, using recombinant inbred strains of mice. More
10 specifically, these studies found that Tnf (protein product is the inflammatory cytokine
11 TNF-a) and Tlr4 (protein product is TLR4, involved in endotoxin responses) were
12 candidate susceptibility genes (Kleeberger et al.. 2000; Kleeberger et al.. 1997). The TNF
13 receptors 1 and 2 have also been found to play a role in injury, inflammation, and airway
14 hyperreactivity in studies of O3-exposed knockout mice (Cho et al.. 2001). In addition to
15 Tlr4, other innate immune pattern recognition signaling pathway genes, including Tlr2
16 and Myd88, appear to be important in responses to O3, as demonstrated by Williams et
17 al. (2007b). A role for the inflammatory cytokine IL-6 has been demonstrated in
18 gene-deficient mice with respect to inflammation and injury, but not AHR (Johnston et
19 al.. 2005b; Yu et al.. 2002). Mice deficient in IL-10, an anti-inflammatory cytokine,
20 demonstrated increased pulmonary inflammation in response to O3 exposure (Backus et
21 al.. 2010). Thus genes related to innate immune signaling and pro- and anti-inflammatory
22 genes are important for O3-induced responses.
23 Altered O3 responses between mouse strains could be due to genetic variability in
24 nuclear factor erythroid 2-related factor 2 (Nrf-2), suggesting a role for genetic
25 differences in altering the formation of ROS (Hamade et al., 2010; Cho and Kleeberger.
26 2007). Additionally, some studies have reported O3-related effects to vary by Inf-1 and
27 Inf-2 quantitative trait loci (Tankersley and Kleeberger. 1994) and a gene coding for
28 Clara cell secretory protein (CCSP) (Broeckaert et al.. 2003; Wattiez et al.. 2003). Other
29 investigations in inbred mouse strains found that differences in expression of certain
30 proteins, such as CCSP (Broeckaert et al.. 2003) and MARCO (Dahl et al.. 2007). are
31 responsible for phenotypic characteristics, such as epithelial permeability and scavenging
32 of oxidized lipids, respectively, which confer sensitivity to O3.
33 Nitric oxide (NO), derived from activated macrophages, is produced upon exposure to O3
34 and is thought to participate in lung damage. Mice deficient in the gene for inducible
35 nitric oxide synthase (NOS2/NOSIMNOS) are partially protected against lung injury
36 (Kleeberger et al.. 2001). and it appears that O3-induced iNOS expression is tied to the
37 TLR4 pathway described above. Similarly, iNOS deficient mice do not produce reactive
38 nitrogen intermediates after O3 exposure, in contrast to their wild-type counterparts, and
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1 also produce less PGE2 comparatively (Fakhrzadeh et al., 2002). These gene-deficient
2 mice were protected from O3-induced lung injury and inflammation. In contrast, another
3 study using a similar exposure concentration but longer duration of exposure found that
4 iNOS deficient mice were more susceptible to O3-induced lung damage (Kenyon et al..
5 2002). Therefore it is unclear whether inducible nitric oxide synthase plays a protective
6 role or mediates damage.
7 Voynow et al. (2009) have shown that NQO1 deficient mice, like their human
8 counterparts, are resistant to O3-induced AHR and inflammation. NQO1 catalyzes the
9 reduction of quinones to hydroquinones, and is capable of both protective detoxification
10 reactions and redox cycling reactions resulting in the generation of reactive oxygen
11 species. Reduced production of inflammatory mediators and cells and blunted AHR were
12 observed in NQO1 null mice after exposure to 1 ppm O3 for 3 hours. These results
13 correlated with those from in vitro experiments in which human bronchial epithelial cells
14 treated with an NQO1 inhibitor exhibited reduced inflammatory responses to exposure to
15 0.4 ppm O3 for 5 hours. This study may provide biological plausibility for the increased
16 biomarkers of oxidative stress and increased pulmonary function decrements observed in
17 O3-exposed individuals bearing both the wild-type NQO1 gene and the null GSTM1 gene
18 (Corradi et al.. 2002: Bergamaschi et al.. 2001).
19 The role of TNF-a signaling in O3-induced responses has been previously established
20 through depletion experiments, but a more recent toxicological study investigated the
21 effects of combined O3 and PM exposure in transgenic TNF overexpressing mice.
22 Kumarathasan et al. (2005) found that subtle effects of these pollutants were difficult to
23 identify in the midst of the severe pathological changes caused by constitutive TNF-a
24 overexpression. However, there was evidence that TNF transgenic mice were more
25 susceptible to O3/PM-induced oxidative stress, and they exhibited elevation of a serum
26 creatine kinase after pollutant exposure, which may suggest potential systemic or cardiac
27 related effects. Differential susceptibility to O3 among inbred strains of animals does not
28 seem to be dose dependent since absorption of 18O in various strains of mice did not
29 correlate with resistance or sensitivity (Vancza et al.. 2009).
30 Defects in DNA repair mechanisms may also confer increased risk of O3-related health
31 effects. Cockayne syndrome, a rare autosomal recessive disorder in humans, is
32 characterized by UV sensitivity abnormalities, neurological abnormalities, and premature
33 aging. The same genetic defect in mice (Csb~A) makes them sensitive to oxidative
34 stressors, including O3. Kooter et al. (2007) demonstrated that Csb"7" mice produced
35 significantly more TNF-a after exposure to 0.8 ppm O3 than their wild-type counterparts.
36 However, there were no significant differences in other markers of inflammation or lung
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1 injury between the two strains of mice. Further discussion of candidate genes in the
2 context of their respective signaling pathways can be found in Chapter 5.
3 Overall, multiple genes, such as GSTM1, GSTP1, HMOX-1, NQO1, and TNF-a, appear
4 to potentially be involved in populations being more at-risk than others to the effects of
5 O3 exposure on health. Future studies of these and other genes in human populations will
6 be important for determining the role of each genotype and its effect on risk. For NQO1
7 and TNF-a, biological plausibility is provided by toxicological studies. Additionally,
8 studies of rodents have identified a number of other genes that may affect O 3 -related
9 health outcomes, but testing of these genes has not been performed in humans due to
10 power limitations.
8.5 Diet
11 Diet was not examined as a factor affecting risk in previous O3 AQCDs, but recent
12 studies have examined modification of the association between O3 and health effects by
13 dietary factors. Because O3 mediates some of its toxic effects through oxidative stress,
14 the antioxidant status of an individual is an important factor that may contribute to
15 increased risk of O3-related health effects. Supplementation with vitamin E has been
16 investigated in a number of studies as a means of inhibiting O3-mediated damage.
17 Epidemiologic studies have examined effect measure modification by diet and found
18 evidence that certain dietary components are related to the effect O3 has on respiratory
19 outcomes. The most recent study examined fruit/vegetable intake and Mediterranean diet
20 (Romieu et al., 2009). Increases in these food patterns, which have been noted for their
21 high vitamins C and E and omega-3 fatty acid content, protected against O3-related
22 decreases in lung function among children living in Mexico City. Another study
23 examined supplementation of the diets of asthmatic children in Mexico with vitamins C
24 and E (Sienra-Monge et al., 2004). Associations were detected between short-term O3
25 exposure and nasal airway inflammation among children in the placebo group but not in
26 those receiving the supplementation. The authors concluded that "vitamin C and E
27 supplementation above the minimum dietary requirement in asthmatic children with a
28 low intake of vitamin E might provide some protection against the nasal acute
29 inflammatory response to ozone."
30 The epidemiologic evidence is supported by controlled human exposure studies, which
31 have shown that the first line of defense against oxidative stress is antioxidants-rich
32 extracellular lining fluid (ELF) which scavenge free radicals and limit lipid peroxidation.
33 Exposure to O3 depletes the antioxidant level in nasal ELF probably due to scrubbing of
34 O3 (Mudway et al.. 1999a): however, the concentration and the activity of antioxidant
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1 enzymes either in ELF or plasma do not appear to be related to O3 responsiveness (Samet
2 etal.. 2001: Avissaretal.. 2000: Blomberg et al.. 1999). Carefully controlled studies of
3 dietary antioxidant supplementation have demonstrated some protective effects of
4 a-tocopherol (a form of vitamin E) and ascorbate (vitamin C) on spirometric measures of
5 lung function after O3 exposure but not on the intensity of subjective symptoms and
6 inflammatory response including cell recruitment, activation and a release of mediators
7 (Samet etal.. 2001: Trengaet al., 2001). Dietary antioxidants have also afforded partial
8 protection to asthmatics by attenuating postexposure bronchial hyperresponsiveness
9 (Trengaetal.. 2001).
10 Toxicological studies provide evidence of biological plausibility to the epidemiologic and
11 controlled human exposure studies. Wagner et al. (2009: 2007) have shown reductions in
12 O 3-exacerbated nasal allergy responses in rats with y-tocopherol treatment (a form of
13 vitamin E). O3-induced inflammation and mucus production were also inhibited by
14 y-tocopherol. Inconsistent results were observed in toxicological studies of vitamin C
15 deficiency and O3-induced responses. Guinea pigs deficient in vitamin C displayed only
16 minimal injury and inflammation after exposure to O3 (Kodavanti et al., 1995). A recent
17 study in mice demonstrated a protective effect of p-carotene in the skin, where it limited
18 the production of proinflammatory markers and indicators of oxidative stress induced by
19 O3 exposure (Valacchi et al., 2009). Deficiency of vitamin A, which has a role in
20 regulating the maintenance and repair of the epithelial layer, particularly in the lung,
21 appears to enhance the risk of O3-induced lung injury (Paquette et al.. 1996).
22 Differentially susceptible strains that were fed a vitamin A sufficient diet were observed
23 to have different tissue concentrations of the vitamin, potentially contributing to their
24 respective differences in O3-related outcomes. In addition to the studies of antioxidants,
25 one toxicological study examined protein deficiency. Protein deficiency alters the levels
26 of enzymes and chemicals in the brain involved with redox status; exposure to 0.75 ppm
27 O3 has been shown to differentially affect Na+/K+ ATPase, glutathione, and lipid
28 peroxidation, depending on the nutritional status of the animal, but the significance of
29 these changes is unclear (Calderon Guzman et al., 2006). There may be a protective
30 effect of overall dietary restriction with respect to lung injury, possibly related to
31 increased vitamin C in the lung surface fluid (Kari etal.. 1997).
32 Epidemiologic studies find that individuals with diets deficient in vitamins E and C are at
33 risk for O3 -related health effects. This is supported by controlled human exposure and
34 toxicological studies.
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8.6 Body Mass Index and Physical Conditioning
1 Obesity, defined as a BMI of 30 kg/m2 or greater, is an issue of increasing importance in
2 the U.S., with self-reported rates of 26.7% in 2009, up from 19.8% in 2000 (Sherry et al..
3 2010). A few studies have been performed examining the association between BMI and
4 O 3-related changes in lung function. An epidemiologic study reported decreased lung
5 function with increased short-term O3 exposure for both obese and non-obese subjects;
6 however, the magnitude of the reduction in lung function was greater for those subjects
7 who were obese (Alexeeff et al., 2007). Further decrements in lung function were noted
8 for obese individuals with AHR. Controlled human exposure studies have also detected
9 differential effects of O3 exposure on lung function for individuals with varying BMIs. In
10 a retrospective analysis of data from 541 healthy, nonsmoking, white males between the
11 ages of 18-35 years from 15 studies conducted at the U.S. EPA Human Studies Facility in
12 Chapel Hill, North Carolina, McDonnell et al. (2010) found that increased body mass
13 index (BMI) was found to be associated with enhanced FEVi responses. The BMI effect
14 was of the same order of magnitude but in the opposite direction of the age effect
15 whereby FEVi responses diminish with increasing age. In a similar analysis, Bennett et
16 al. (2007) found enhanced FEVi decrements following O3 exposure with increasing BMI
17 in a group of healthy, nonsmoking, women (BMI range 15.7 to 33.4), but not among
18 healthy, nonsmoking men (BMI range 19.1 to 32.9). In the women, greater O3-induced
19 FEVi decrements were seen in individuals that were overweight/obese (BMI >25)
20 compared normal weight (BMI from 18.5 to 25), and in normal weight compared to
21 underweight (BMI <18.5). Even disregarding the five underweight women, a greater O3
22 response in the overweight/obese category (BMI >25) was observed compared with the
23 normal weight group (BMI from 18.5 to 24.9).
24 Studies in genetically and dietarily obese mice have shown enhanced pulmonary
25 inflammation and injury with acute O3exposure, but responses to longer exposures at a
26 lower concentration appear to differ. A recent study found that obese mice are actually
27 resistant to O3-induced pulmonary injury and inflammation and reduced lung compliance
28 following exposure to 0.3 ppm O3 for 72 hours, regardless of whether obesity was
29 genetic- or diet-induced (Shore et al.. 2009).
30 In addition to studies of obesity, physical conditioning affects BMI and may also affect
31 the risk of O3-related health effects. The 2008 Summary of Health Statistics for U.S.
32 Adults from the CDC reported the prevalence of regular leisure-time physical activity as
33 slightly above 30% for adults >18 years of age in the U.S. (Pleis et al.. 2009). Forty-nine
34 percent of individuals >65 years old reported no leisure-time physical activity. A study of
3 5 effect measure modification by exercise habits ten years prior to death observed excess
36 risk of mortality with increasing O3 concentrations among individuals that never
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1 exercised compared to individuals that exercised at least once a month for both adults
2 >30 years of age and adults >65 years of age (Wong et al.. 2007). No recent studies
3 examining modification of O3 -related health effects by current physical activity were
4 identified.
5 Multiple epidemiologic and human clinical studies have reported increased O3-related
6 respiratory health effects among obese individuals. Future research of the effect
7 modification of the relationship between O3 and other health-related outcomes besides
8 respiratory health effects by BMI and studies examining the role of physical conditioning
9 will advance understanding of obesity as a factor potentially increasing an individual's
10 risk.
8.7 Socioeconomic Status
11 SES is often represented by personal or neighborhood SES, educational attainment,
12 health insurance status, and other such factors. SES is indicative of such things as access
13 to healthcare, quality of housing, and pollution gradient. Based on the 2000 Census data,
14 12.4% of Americans live in poverty (poverty threshold for family of four was $17,463)
15 (SSDAN CensusScope. 2010c).
16 Multiple epidemiologic studies have reported individuals of low SES to have increased
17 risk for the effects of short-term O3 exposure on respiratory HAs and ED visits. A study
18 performed in Korea examined the association between O3 concentrations and asthma HA
19 and reported larger effect estimates in areas of moderate and low SES compared with
20 areas of high SES (SES was based on average regional insurance rates) (Lee et al.. 2006).
21 A Canadian study reported inverse effects of O3 on respiratory HA and ED visits
22 regardless of SES, measured by average census tract household income (Burra et al..
23 2009). In addition, a study conducted across 10 cities in Canada found the largest
24 association between O3 exposure and respiratory HA was among those with an
25 educational level less than grade 9, but no consistent trend in the effect was seen across
26 quartiles of income (Cakmak et al.. 2006b). In New York State, larger associations
27 between long-term O3 exposure and asthma HA were observed among children of
28 mothers who did not graduate from high school, whose births were covered by
29 Medicaid/self-paid, or who were living in poor neighborhoods compared to children
30 whose mothers graduated from high school, whose births were covered by other
31 insurance, or who were not living in poor neighborhoods, respectively (Lin et al.. 2008b).
32 The examination of the potential effects of SES on O3-related cardiovascular health
33 effects is relatively limited. A study conducted in Canada reported the association
34 between short-term O3 and ED visits for cardiac disease by quartiles of
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1 neighborhood-level education and income. No effect measure modification was apparent
2 for either measure of SES (Cakmak et al.. 2006a).
3 Several studies were conducted that examined the modification of the relationship
4 between short-term O3 concentrations and mortality by SES. A U.S. multicity study
5 reported that communities with a higher proportion of the population unemployed had
6 higher mortality effect estimates (Bell and Dominici. 2008). A study in seven urban
7 centers in Chile reported on modification of the association between O3 exposure and
8 mortality using multiple SES markers (Cakmak et al.. 2011). Increased risk was observed
9 among the categories of low SES for all measures (personal educational attainment,
10 personal occupation, community income level). Additionally, the APHENA study, which
11 examined the association between O3 and mortality by percentage unemployed, reported
12 a higher percent change in mortality with increased percent unemployed but this varied
13 across the regions included in the study (U.S., Canada, Europe) (Katsouyanni et al..
14 2009). A Chinese study reported that the greatest effects between O3 concentrations and
15 mortality at lag day 0 were among individuals living in areas of high social deprivation
16 (i.e. low SES), but this association was not consistent across lag days (at other lag times,
17 the middle social deprivation index category had the greatest association) (Wong et al..
18 2008). However, another study in Asia comparing low to high educational attainment
19 populations reported no evidence of greater mortality effects (total, CVD, or respiratory)
20 (Kan et al.. 2008). Additionally, a study in Italy reported no difference in risk of mortality
21 among census-block level derived income levels (Stafoggia et al.. 2010). A study of
22 infant mortality in Mexico reported no association between O3 concentrations and infant
23 mortality among any of the three levels of SES determined using a socioeconomic index
24 based on residential areas (Romieu et al.. 2004b). Another study in Mexico reported a
25 positive association between O3 levels at lag 0 and respiratory-related infant mortality in
26 only the low SES group (determined based on education, income, and household
27 conditions across residential areas), but no association was observed in any of the SES
28 groups with other lags (Carbajal-Arroyo et al.. 2011).
29 Studies of O3 concentrations and reproductive outcomes have also examined associations
30 by SES levels. A study in California reported greater decreases in birth weight associated
31 with full pregnancy O3 concentration for those with neighborhood poverty levels of at
32 least 7% compared with those in neighborhoods with less than 7% poverty (Morello-
33 Frosch et al.. 2010). However, no dose response was apparent and those with
34 neighborhood poverty levels of 7-21% had greater decreases observed for the association
3 5 than those living in areas with poverty rates of at least 22%. An Australian study reported
36 an inverse association between O3 exposure during days 31-60 of gestation and
37 abdominal circumference during gestation (Hansen et al.. 2008). The interaction with
3 8 SES (area-level measured socioeconomic disadvantage) was examined and although the
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1 inverse association remained statistically significant in only the highest SES quartile,
2 there were large confidence interval overlaps among estimates for each quartile so no
3 difference in the association for the quartiles was apparent.
4 Evidence from a controlled human exposure study that examined O3 effects on lung
5 function does not provide support for greater O3-related health effects in individuals of
6 lower SES. In a follow-up study (Seal et al., 1993) on modification by race, Seal et al.
7 (1996) reported that, of three SES categories, individuals in the middle SES category
8 showed greater concentration-dependent decline in percent-predicted FEVi (4-5% at
9 400 ppb O3) than in low and high SES groups. The authors did not have an "immediately
10 clear" explanation for this finding and controlled human exposure studies are typically
11 not designed to answer questions about SES.
12 Overall, most studies of individuals have reported that individuals with low SES and
13 those living in neighborhoods with low SES are more at risk for O3-related health effects
14 resulting in higher odds of respiratory HAs and ED visits. Inconsistent results have been
15 observed in the few studies examining effect modification of associations between O3
16 exposure and mortality and reproductive outcomes.
8.8 Race/Ethnicity
17 Based on the 2000 Census, 69.1% of the U.S. population comprises non-Hispanic whites.
18 Approximately 12.1% of people reported their race/ethnicity as non-Hispanic black and
19 12.6% reported being Hispanic (SSDAN CensusScope. 201 Ob).
20 Two studies examined the associations between short-term O3 concentrations and
21 mortality and reported higher effect estimates among blacks (Medina-Ramon and
22 Schwartz. 2008) and among communities with larger proportions of blacks (Bell and
23 Dominici. 2008). Another study examined long-term exposure to O3 concentrations and
24 asthma HAs among children in New York State. These authors reported no statistically
25 significant difference in the odds of asthma HA for blacks compared to other races but
26 did detect higher odds for Hispanics compared to non-Hispanics (Lin et al.. 2008b).
27 Additionally, recent epidemiologic studies have stratified by race when examining the
28 association between O3 concentration and birth outcomes. A study conducted in Atlanta,
29 GA reported decreases in birth weight with increased third trimester O3 concentrations
30 among Hispanics but not among non-Hispanic whites (Darrow et al., 201 la). An inverse
31 association was also present for non-Hispanic blacks but was not statistically significant.
32 A California study reported that the greatest decrease in birth weight associated with full
33 pregnancy O3 concentration was among non-Hispanic whites (Morello-Frosch et al..
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1 2010). The inverse association was also apparent, although not as strong, for
2 non-Hispanic blacks. Increased birth weight was associated with higher O3 exposure
3 among Hispanics and among non-Hispanic Asians and Pacific Islanders but neither of
4 these results were statistically significant.
5 Similar to the epidemiologic studies, a controlled human exposure study suggested
6 differences in lung function responses by race (Seal et al., 1993). The independent effects
7 of sex-race group and O3 concentration on lung function were positive, but the
8 interaction between sex-race group and O3 concentration was not statistically significant.
9 The findings indicated some overall difference between the sex-race groups that was
10 independent of O3 concentration (the concentration-response curves for the four sex-race
11 groups are parallel). In a multiple comparison procedure on data collapsed across all O3
12 concentrations for each sex-race group, both black men and black women had larger
13 decrements in FEVi than did white men. The authors noted that the O3 dose per unit of
14 lung tissue would be greater in blacks and females than whites and males, respectively.
15 That this difference in tissue dose might have affected responses to O3 cannot be ruled
16 out. The college students recruited for the Seal et al. (1993) study were probably from
17 belter educated and more SES advantaged families, thus reducing potential for these
18 variables to be confounding factors. Que et al. also examined pulmonary responses to O3
19 exposure in blacks of African American ancestry and in whites. On average, the black
20 males experienced the greatest decrements in FEVi following O3 exposure. This
21 decrease was larger than the decrement observed among black females, white males, and
22 white females.
23 Overall, the results of recent studies suggest that there may be race-related increase in
24 risk of O3-related health effects for some outcomes, although the overall understanding of
25 potential effect measure modification by race is limited by the small number of studies.
26 Additionally, these results may be confounded by other factors, such as SES.
8.9 Smoking
27 Previous O3 AQCDs have concluded that smoking does not increase the risk of
28 O3-related health effects; in fact, in controlled human exposure studies, smokers have
29 been found to be at less risk of O3-related health effects than non-smokers. Data from
30 recent interviews conducted as part of the 2008 National Health Interview Survey (NHIS)
31 (Pleis et al.. 2009) have shown the rate of smoking among adults >18 years old to be
32 approximately 20% in the U.S. Approximately 21% of individuals surveyed were
33 identified as former smokers.
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1 Baccarelli et al. (2007) performed a study of O3 concentrations and plasma homocysteine
2 levels (a risk factor for vascular disease). They found no interaction of smoking (smokers
3 versus non-smokers) for the associations between O3 concentrations and plasma
4 homocysteine levels. Another study examined the association between O3 and resting
5 heart rate and also reported no interaction with smoking status (current smokers versus
6 current non-smokers) (Ruidavets et al.. 2005a).
7 A study examining correlations between O3 levels and oxidative DNA damage examined
8 results stratified by current versus never and former smokers (Palli et al.. 2009). Ozone
9 was positively associated with DNA damage for short-term and long-term exposures
10 among never/former smokers. For current smokers, short-term O3 concentrations were
11 inversely associated with DNA damage; however, the number of current smokers in the
12 study was small (n= 12).
13 The findings of Palli et al. (2009) were consistent with those from controlled human
14 exposure studies that have confirmed that smokers are less responsive to O3 exposure
15 than non-smokers. Spirometric and plethysmographic pulmonary function decline,
16 nonspecific AHR, and inflammatory responses of smokers to O3 exposure were all
17 weaker than those reported for non-smokers. Similarly, the time course of development
18 and recovery from these effects, as well as their reproducibility, was not different from
19 non-smokers. Chronic airway inflammation with desensitization of bronchial nerve
20 endings and an increased production of mucus may plausibly explain the
21 pseudo-protective effect of smoking (Frampton et al.. 1997b: Torres et al.. 1997).
22 These findings for smoking are consistent with previous AQCD conclusions. An
23 epidemiologic study of O3-associated DNA damage reported smokers to be less at risk
24 for O3-related health effects. However, both epidemiologic studies of short-term
25 exposure and CVD outcomes found no effect measure modification by smoking.
8.10 Heightened Exposure
26 Studies included in the 2006 O3 AQCD reported that individuals who participate in
27 outdoor activities or work to be a population at increased risk based on consistently
28 reported associations between O3 exposure and respiratory health outcomes in these
29 groups (U.S. EPA. 2006b). Outdoor workers are exposed to ambient O3 concentrations
30 outside for a greater period of time than individuals who spend their days indoors.
31 Additionally, an increase in dose to the lower airways is possible during exercise due to
32 both increases in the amount of air breathed (i.e., minute ventilation) and a shift from
33 nasal to oronasal breathing (Sawyer et al., 2007; Nodelman and Ultman, 1999; Hu et al.,
34 1994). For further discussion of the association between FEVi responses to O3 exposure
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1 and minute ventilation, refer to Section 6.2.3.1 of the 2006 O3 AQCD. A recent study has
2 explored the potential effect measure modification of O3 exposure and DNA damage by
3 indoor/outdoor workplace (Tovalin et al., 2006). In a study of indoor and outdoor
4 workers in Mexico, individuals who worked outdoors in Mexico City had a slight
5 association between O3 exposure and DNA damage (measured by comet tail length
6 assay), whereas no association was observed for indoor workers in Mexico City. Workers
7 in another Mexican city, Puebla, demonstrated no association between O3 levels and
8 DNA damage, regardless of whether they worked indoors or outdoors.
9 Air conditioning use is an important component of O3 exposure, as use of central air
10 conditioning will limit exposure to O3 by blocking the penetration of O3 into the indoor
11 environment (further information can be found in Section 4.4 of this ISA). Air
12 conditioning use is a difficult effect measure modifier to examine in epidemiologic
13 studies. Air conditioning use is often measured based on regional prevalence and may not
14 reflect individua!4evel use. More generally, air conditioning prevalence is associated
15 with temperature of a region; those areas with higher temperatures have a greater
16 prevalence of households with air conditioning. Despite these limitations, a few studies
17 have examined effect measure modification by prevalence of air conditioning use in an
18 area. Studies examining multiple cities across the U.S. have assessed whether
19 associations between O3 concentrations and HA and mortality varied among areas with
20 high and low prevalence of air conditioning. Medina-Ramon et al. (2006) conducted a
21 study during the warm season and observed a greater association between O3 levels and
22 pneumonia HA among areas with a lower proportion of households having central air
23 conditioning compared to areas with a larger proportion of households with air
24 conditioning. The same trend of increased association for areas with a lower prevalence
25 of central air conditioning was noted in a study of O3 concentrations and mortality (Bell
26 and Dominici. 2008). Conversely, Medina-Ramon and Schwartz (2008) found that
27 among individuals with atrial fibrillation, a lower risk of mortality was observed for areas
28 with a lower prevalence of central air conditioning.
29 Previous work has shown that increased dose of O3 concentrations from outdoor work
30 leads to increased risk of O3-related health effects among individuals who participate in
31 outdoor activities or work, although there is no evidence of modification by outdoor
32 activity in this recent study. Lower prevalence of air conditioning also appears to affect
33 risk of O3-related health effects, but this is not true of all studies. Overall, increased
34 exposure to outdoor air does appear to confer additional risk and individuals with greater
35 exposure to outdoor air may experience more O3-related health effects.
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8.11 Healthy Responders
1 Within the general population, there is evidence for variability in responses to O3
2 exposure, with some healthy individuals demonstrating greater O3-related health effects
3 compared to other healthy individuals in controlled human exposure studies. These
4 individuals do not fit in any of the at-risk populations discussed in this chapter; however,
5 studies have found that they have greater responses to O3 exposure than would be
6 expected, indicating a unique population that needs to be considered.
7 Controlled human exposure studies have demonstrated a large degree of intersubject
8 variability in lung function decrements, symptomatic responses, pulmonary
9 inflammation, AHR, and altered epithelial permeability in healthy adults exposed to O3
10 (Que et al.; Holz et al., 2005; McDonnell. 1996). The magnitude of increases in
11 pulmonary inflammation, AHR, and epithelial permeability, in response to O3 exposure,
12 do not appear to be correlated, nor are these responses correlated with changes in lung
13 function (Que et al.: Balmesetal.. 1997; Balmesetal.. 1996; Arisetal.. 1995). However,
14 these responses to O3 exposure in healthy individuals tend to be reproducible within a
15 given individual over a period of several months indicating differences in the intrinsic
16 responsiveness of individuals (Holz et al.. 2005; Hazucha et al.. 2003; Holz et al.. 1999;
17 McDonnell et al.. 1985a). It should be noted that even when group mean responses are
18 small and seem physiologically insignificant, some intrinsically more responsive
19 individuals experience distinctly larger effects under the same exposure conditions. For
20 example, small group mean changes (e.g., <5%) in FEVi have been observed in healthy
21 young adults at levels as low as 120 ppb O3 for 1 to 3 hour exposure periods. However,
22 some individuals within a study may experience FEVi decrements in excess of 15%
23 under these conditions, even with group mean decrements of less than 5%. Therefore,
24 within the general population, a proportion of otherwise healthy individuals, who do not
25 have characteristics discussed above that increase risk, may be at increased risk of
26 O 3 -induced health effects.
8.12 Summary
27 In this section, epidemiologic, controlled human exposure, and toxicological studies have
28 been evaluated that contribute information on potential at-risk populations. Overall, this
29 review provides evidence that various factors may lead to increased risk of O3-related
30 health effects.
31 The populations identified in this section that are most at risk for O3-related health effects
32 are individuals with influenza/infection, individuals with asthma, and younger and older
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1 age groups. There were a small number of studies on influenza/infection but both
2 reported influenza/infection to modify the association between O3 exposure and
3 respiratory effects, with individuals having influenza or an infection being at increased
4 risk. Asthma as a factor affecting risk was supported by controlled human exposure and
5 toxicological studies, as well as some evidence from epidemiologic studies. Most studies
6 comparing age groups reported greater effects of short-term O3 exposure on mortality
7 among older adults, although studies of other health outcomes had inconsistent findings
8 regarding whether older adults were at increased risk. Generally, studies of age groups
9 also reported positive associations for respiratory HAs and ED visits among children.
10 Biological plausibility for this increased risk is supported by toxicological and clinical
11 research. Diet and obesity are also both likely factors affecting risk. Multiple
12 epidemiologic, controlled human exposure, and toxicological studies reported that diets
13 deficient in vitamins E and C are associated with risk of O3-related health effects.
14 Similarly, studies of effect measure modification by BMI observed greater O3-related
15 respiratory decrements for individuals who were obese.
16 Other potential factors [preexisting conditions (such as COPD and CVD), sex, and
17 multiple genes (such as GSJM1, GSTP1, HMOX-1, NQO1, and TNF-a)} provided some
18 evidence of increased risk, but further evidence is needed. In addition, examination of
19 modification of the associations between O3 exposure and health effects by SES and race
20 were available in a limited number of studies, and demonstrated possible increased odds
21 of health effects related to O3 exposure among those with low SES and black race.
22 Individuals with increased outdoor exposure were examined in a recent study of outdoor
23 workers, in which no effect modification was observed, and studies of air conditioning
24 prevalence, which demonstrated inconsistent findings. However, previous evidence along
25 with biological plausibility from toxicological and controlled human studies has shown
26 individuals exposed to more outdoor air to be at increased risk of O3-related health
27 effects. Studies of physical conditioning and smoking were conducted but little evidence
28 was available to determine whether increased risk of O3-related health effects is present
29 for these factors. The only studies examining effect measure modification by diabetes
30 examined O3 exposure and cardiovascular outcomes and none reported increased risks for
31 individuals with diabetes. Toxicological studies also identified hyperthyroidism to be a
32 factor warranting further examination. Future research will provide additional insight into
33 whether these factors affect risk of O3-related health effects.
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9 ENVIRONMENTAL EFFECTS: OZONE EFFECTS
ON VEGETATION AND ECOSYSTEMS
9.1 Introduction
1 This chapter synthesizes and evaluates the relevant science to help form the scientific
2 foundation for the review of a vegetation- and ecologically-based secondary NAAQS for
3 O3. The secondary NAAQS are based on welfare effects. The Clean Air Act (CAA)
4 definition of welfare effects includes, but is not limited to, effects on soils, water,
5 wildlife, vegetation, visibility, weather, and climate, as well as effects on materials,
6 economic values, and personal comfort and well-being. The effects of O3 as a greenhouse
7 gas and its direct effects on climate are discussed in Chapter 10 of this document.
8 The intent of the ISA, according to the CAA, is to "accurately reflect the latest scientific
9 knowledge expected from the presence of [a] pollutant in ambient air" (42 U.S.C.7408
10 and 42 U.S.C.7409 (1999). This chapter of the ISA includes scientific research from
11 biogeochemistry, soil science, plant physiology, and ecology conducted at multiple scales
12 (e-g-, organ, organism, population, community, ecosystem). Key information and
13 judgments formerly found in the AQCDs regarding O3 effects on vegetation and
14 ecosystems are found in this chapter. This chapter of the O3 ISA serves to update and
15 revise Chapter 9 and AX9 of the 2006 O3 AQCD (U.S. EPA. 2006b).
16 Numerous studies of the effects of O3 on vegetation and ecosystems were reviewed in the
17 2006 O3 AQCD. That document concluded that the effects of ambient O3 on vegetation
18 and ecosystems appear to be widespread across the U.S., and experimental studies
19 demonstrated plausible mechanisms for these effects. Ozone effect studies published
20 from 2005 to July 2011 are reviewed in this document in the context of the previous O3
21 AQCDs. From 2005 to 2011, some areas have had very little new research published and
22 the reader is referred back to sections of the 2006 O3 AQCD for a more comprehensive
23 discussion of those subjects. This chapter is focused on studies of vegetation and
24 ecosystems that occur in the U.S. and that report endpoints or processes most relevant to
25 the review of the secondary standard. Many studies have been published about vegetation
26 and ecosystems outside of the U.S. and North America, largely in Europe and Asia. This
27 document includes discussion of studies of vegetation and ecosystems outside of North
28 America only if those studies contribute to the general understanding of O3 effects across
29 species and ecosystems. For example, studies outside North America are discussed that
30 consider physiological and biochemical processes that contribute to the understanding of
31 effects of O3 across species. Also, ecosystem studies outside of North America that
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1 contribute to the understanding of O3 effects on general ecosystem processes are
2 discussed in the chapter.
3 Sections of this chapter first discuss exposure methods, followed by effects on vegetation
4 and ecosystems at various spatial scales and ends with policy-relevant discussions of
5 exposure indices and exposure-response. Figure 9-1 is a simplified illustrative diagram of
6 the major pathway through which O3 enters plants and the major endpoints O3 may
7 affect. First, Section 9.2 presents a brief overview of various methodologies that have
8 been, and continue to be, central to quantifying O3 effects on vegetation (AX9.1 of the
9 2006 O3 AQCD for more detailed discussion) (U.S. EPA. 2006b). Sections 9.3 through
10 9.4 begin with a discussion of effects at the cellular and subcellular level followed by
11 consideration of the O3 effects on plant and ecosystem processes (Figure 9-1). In Section
12 9.3, research is reviewed from the molecular to the biochemical and physiological levels
13 in impacted plants, offering insight into the mode of action of O3. Section 9.4 provides a
14 review of the effects of O3 exposure on major endpoints at the whole plant scale
15 including growth, reproduction, visible foliar injury and leaf gas exchange in woody and
16 herbaceous plants in the U.S., as well as a brief discussion of O3 effects on agricultural
17 crop yield and quality. Section 9.4 also integrates the effects of O3 on individual plants in
18 a discussion of available research for assessing the effect of O3 on ecosystems, along
19 with available studies that could inform assessments of various ecosystem services (See
20 section 9.4.1.2). The development of indices of O3 exposure and dose modeling is
21 discussed in Section 9.5. Finally, exposure-response relationships for a number of tree
22 species, native vegetation, and crop species and cultivars are reviewed, tabulated, and
23 compared in Section 9.6 to form the basis for an assessment of the potential risk to
24 vegetation from current ambient levels of O3.
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03 exposure
03 uptake & physiology (Fig 9-2)
•Antioxidant metabolism up-regulated
•Decreased photosynthesis
•Decreased stomatal conductance
or sluggish stomatal response
Effects on leaves
Visible leaf injury
Altered leaf production
Altered leaf chemical composition
Plant growth (Fig 9.8)
•Decreased biomass accumulation
•Altered reproduction
•Altered carbon allocation
•Altered crop quality
Affected ecosystem services
•Decreased productivity
•Decreased C sequestration
•Altered water cycling (Fig 9-7)
•Altered community composition
(i.e., plant, insects microbe)
Belowground processes (Fig 9.8)
•Altered litter production and decomposition
•Altered soil carbon and nutrient cycling
•Altered soil fauna and microbial communities
Figure 9-1 An illustrative diagram of the major pathway through which Oz
enters plants and the major endpoints that Os may affect in plants
and ecosystems.
9.2 Experimental Exposure Methodologies
9.2.1 Introduction
1
2
3
4
5
6
7
A variety of methods for studying plant response to O3 exposures have been developed
over the last several decades. Methodological advancements since 2006 have not
fundamentally altered our understanding of O3 effects on plants or ecosystems. The
majority of methodologies currently used have been discussed in detail in the 1996 O3
AQCD and 2006 O3 AQCD. This section will serve as a short overview of the
methodologies and the reader is referred to the previous O3 AQCDs for more in-depth
discussion.
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9.2.2 "Indoor," Controlled Environment, and Greenhouse Chambers
1 The earliest experimental investigations of the effects of O3 on plants utilized simple
2 glass or plastic-covered chambers, often located within greenhouses, into which a flow of
3 O3-enriched air or oxygen could be passed to provide the exposure. The types, shapes,
4 styles, materials of construction, and locations of these chambers have been numerous.
5 Hogsett et al. (1987a) have summarized the construction and performance of more
6 elaborate and better instrumented chambers since the 1960s, including those installed in
7 greenhouses (with or without some control of temperature and light intensity).
8 One greenhouse chamber approach that continues to yield useful information on the
9 relationships of O3 uptake to both physiological and growth effects employs continuous
10 stirred tank reactors (CSTRs) first described by Heck et al. (1978). Although originally
11 developed to permit mass-balance studies of O3 flux to plants, their use has more recently
12 widened to include short-term physiological and growth studies of O3 * CO2 interactions
13 (Loats and Rebbeck. 1999; Reinert et al.. 1997; Raoetal.. 1995; Reinert and Ho. 1995;
14 Heagle etal.. 1994a). and validation of visible foliar injury on a variety of plant species
15 (Kline et al., 2009; Orendovici et al., 2003). In many cases, supplementary lighting and
16 temperature control of the surrounding structure have been used to control or modify the
17 environmental conditions (Heagle et al., 1994a).
18 Many investigations have utilized commercially available controlled environment
19 chambers and walk-in rooms adapted to permit the introduction of a flow of O3 into the
20 controlled air-volume. Such chambers continue to find use in genetic screening and in
21 physiological and biochemical studies aimed primarily at improving our understanding of
22 modes of action. For example, some of the studies of the O3 responses of common
23 plantain (Plantago major) populations have been conducted in controlled environment
24 chambers (Whitfield et al.. 1996: Reiling and Davison. 1994).
25 More recently, some researchers have been interested in attempting to investigate direct
26 O3 effects on reproductive processes, separate from the effects on vegetative processes
27 (Black etal.. 2010). For this purpose, controlled exposure systems have been employed
28 to expose the reproductive structures of annual plants to gaseous pollutants independently
29 of the vegetative component (Black etal.. 2010; Stewart et al.. 1996).
9.2.3 Field Chambers
30 In general, field chamber studies are dominated by the use of various versions of the open
31 top chamber (OTC) design, first described by Heagle et al. (1973) and Mandl et al.
32 (1973). The OTC method continues to be a widely used technique in the U.S. and Europe
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1 for exposing plants to varying levels of O3. Most of the new information confirms earlier
2 conclusions and provides additional support for OTC use in assessing plant species and in
3 developing exposure-response relationships. Chambers are generally ~3 m in diameter
4 with 2.5-m-high walls. Hogsett et al. (1987b) described in detail many of the various
5 modifications to the original OTC designs that appeared subsequently, e.g., the use of
6 larger chambers for exposing small trees (Kats etal.. 1985) or grapevines (Mandl et al..
7 1989). the addition of a conical baffle at the top to improve ventilation (Kats etal.. 1976).
8 a frustum at the top to reduce ambient air incursions, and a plastic rain-cap to exclude
9 precipitation (Hogsett et al.. 1985). All versions of OTCs included the discharge of air via
10 ports in annular ducting or interiorly perforated double-layered walls at the base of the
11 chambers to provide turbulent mixing and the upward mass flow of air.
12 Chambered systems, including OTCs, have several advantages. For instance, they can
13 provide a range of treatment levels including charcoal-filtered (CF), clean-air control, and
14 several above ambient concentrations for O3 experiments. Depending on experimental
15 intent, a replicated, clean-air control treatment is an essential component in many
16 experimental designs. The OTC can provide a consistent, definable exposure because of
17 the constant wind speed and delivery systems. Statistically robust concentration-response
18 (C-R) functions can be developed using such systems for evaluating the implications of
19 various alternative air quality scenarios on vegetation response. Nonetheless, there are
20 several characteristics of the OTC design and operation that can lead to exposures that
21 might differ from those experienced by plants in the field. First, the OTC plants are
22 subjected to constant air flow turbulence, which, by lowering the boundary layer
23 resistance to diffusion, may result in increased uptake. This may lead to an
24 overestimation of effects relative to areas with less turbulence (Krupaet al.. 1995; Legge
25 etal.. 1995). However, other research has found that OTC's may slightly change vapor
26 pressure deficit (VPD) in a way that may decrease the uptake of O3 into leaves (Piikki et
27 al.. 2008b). As with all methods that expose vegetation to modified O3 concentrations in
28 chambers, OTCs create internal environments that differ from ambient air. This so-called
29 "chamber effect" refers to the modification of microclimatic variables, including reduced
30 and uneven light intensity, uneven rainfall, constant wind speed, reduced dew formation,
31 and increased air temperatures (Fuhrer. 1994; Manning and Krupa. 1992). However, in at
32 least one case where canopy resistance was quantified in OTCs and in the field, it was
33 determined that gaseous pollutant exposure to crops in OTCs was similar to that which
34 would have occurred at the same concentration in the field (Unsworth et al.. 1984a. b).
35 Because of the standardized methodology and protocols used in National Crop Loss
36 Assessment Network (NCLAN) and other programs, the database can be assumed to be
37 internally consistent.
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1 While it is clear that OTCs can alter some aspects of the microenvironment and plant
2 growth, it is important to establish whether or not these differences affect the relative
3 response of a plant to O3. As noted in the 1996 O3 AQCD, evidence from a number of
4 comparative studies of OTCs and other exposure systems suggested that responses were
5 essentially the same regardless of exposure system used and chamber effects did not
6 significantly affect response. For example, a study of chamber effects examined the
7 responses of tolerant and sensitive white clover clones (Trifolium repens) to ambient O3
8 in greenhouse, open top, and ambient plots (Heagle etal.. 1996). The response found in
9 OTCs was the same as in ambient plots.
10 Another type of field chamber called a "terracosm" has been developed and used in
11 recent studies (Lee et al.. 2009a). Concern over the need to establish realistic plant-litter-
12 soil relationships as a prerequisite to studies of the effects of O3 and CO2 enrichment on
13 ponderosa pine (Pinusponderosa) seedlings led Tingey et al. (1996) to develop closed,
14 partially environmentally controlled, sun-lit chambers ("terracosms") incorporating 1-m-
15 deep lysimeters containing forest soil in which the appropriate horizon structure was
16 retained.
17 Other researchers have recently published studies using another type of out-door chamber
18 called recirculating Outdoor Plant Environment Chambers (OPECs) (Flowers et al..
19 2007). These closed chambers are approximately 2.44 mx 1.52 m with a growth volume
20 of approximately 3.7 m3 in each chamber. These chambers admit 90% of full sunlight and
21 control temperature, humidity and vapor pressure (Tiscus etal.. 1999).
9.2.4 Plume and FACE-Type Systems
22 Plume systems are chamberless exposure facilities in which the atmosphere surrounding
23 plants in the field is modified by the injection of pollutant gas into the air above or
24 around them from multiple orifices spaced to permit diffusion and turbulence, so as to
25 establish relatively homogeneous conditions as the individual plumes disperse and mix
26 with the ambient air. They can only be used to increase the O3 levels in the ambient air.
27 The most common plume system used in the U.S. is a modification of the free-air carbon-
28 dioxide/ozone enrichment (FACE) system (Hendrey et al., 1999; Hendrey and Kimball.
29 1994). Although originally designed to provide chamberless field facilities for studying
30 the CO2 effects of climate change, FACE systems have been adapted to include the
31 dispensing of O3 (Karnosky et al.. 1999). This method has been employed in Illinois
32 (SoyFACE) to study soybeans (Morgan et al., 2004; Rogers et al., 2004) and in
33 Wisconsin (Aspen FACE) to study trembling aspen (Populus tremuloides), birch (Betula
34 papyriferd) and maple (Acer saccharum) (Karnosky et al.. 1999). Volk et al. (2003) also
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1 described a similar system for exposing grasslands that uses 7-m diameter plots. FACE
2 systems discharge the pollutant gas (O3 and/or CO2) through orifices spaced along an
3 annular ring (or torus) or at different heights on a ring of vertical pipes. Computer-
4 controlled feedback from the monitoring of gas concentration regulates the feed rate of
5 enriched air to the dispersion pipes. Feedback of wind speed and direction information
6 ensures that the discharges only occur upwind of the treatment plots, and that discharge is
7 restricted or closed down during periods of low wind speed or calm conditions. The
8 diameter of the arrays and their height (25-30 m) in some FACE systems requires large
9 throughputs of enriched air per plot, particularly in forest tree systems. The cost of the
10 throughputs tends to limit the number of enrichment treatments, although Hendrey et al.
11 (1999) argued that the cost on an enriched volume basis is comparable to that of chamber
12 systems.
13 Although plume systems make virtually none of the modifications to the physical
14 environment that are inevitable with chambers, their successful use depends on selecting
15 the appropriate numbers, sizes, and orientations of the discharge orifices to avoid "hot-
16 spots" resulting from the direct impingement of jets of pollutant-enriched air on plant
17 foliage (Werner and Fabian. 2002). Because mixing is unassisted and completely
18 dependent on wind turbulence and diffusion, local gradients are inevitable especially in
19 large-scale systems. FACE systems have provisions for shutting down under low wind
20 speed or calm conditions and for an experimental area that is usually defined within a
21 generous border in order to strive for homogeneity of the exposure concentrations within
22 the treatment area. They are also dependent upon continuous computer-controlled
23 feedback of the O3 concentrations in the mixed treated air and of the meteorological
24 conditions. Plume and FACE systems also are unable to reduce O3 levels below ambient
25 in areas where O3 concentrations are phytotoxic.
9.2.5 Ambient Gradients
26 Ambient O3 gradients that occur in the U.S. hold potential for the examination of plant
27 responses over multiple levels of exposure. However, few such gradients can be found
28 that meet the rigorous statistical requirements for comparable site characteristics such as
29 soil type, temperature, rainfall, radiation, and aspect (Manning and Krupa. 1992):
30 although with small plants, soil variability can be avoided by the use of plants in large
31 pots. The use of soil monoliths transported to various locations along natural O3 gradients
32 is another possible approach to overcome differences in soils; however, this approach is
33 also limited to small plants.
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1 Studies in the 1970s used the natural gradients occurring in southern California to assess
2 yield losses of alfalfa and tomato (Oshima et al.. 1977; Oshima et al.. 1976). A transect
3 study of the impact of O3 on the growth of white clover and barley in the U.K. was
4 confounded by differences in the concurrent gradients of SO2 and NO2 pollution
5 (Ashmore etal.. 1988). Studies of forest tree species in national parks in the eastern U.S.
6 (Winner et al.. 1989) revealed increasing gradients of O3 and visible foliar injury with
7 increased elevation.
8 Several studies have used the San Bernardino Mountains Gradient Study in southern
9 California to study the effects of O3 and N deposition on forests dominated by ponderosa
10 and Jeffrey pine (Jones and Paine. 2006; Arbaugh et al.. 2003; Grulke. 1999; Miller and
11 Elderman. 1977). However, it is difficult to separate the effects of N and O3 in some
12 instances in these studies (Arbaugh et al., 2003). An O3 gradient in Wisconsin has been
13 used to study foliar injury in a series of trembling aspen clones (Populus tremuloides)
14 differing in O3 sensitivity (Mankovska et al., 2005; Karnosky et al.. 1999).
15 More recently, studies have been published that have used natural gradients to study a
16 variety of endpoints and species. For example, Gregg et al. (2003) studied cottonwood
17 saplings grown in an urban to rural gradient of O3 in the New York City area. The
18 secondary nature of the reactions of O3 formation and NOX titration reactions within the
19 city center resulted in significantly higher cumulative O3 exposures in the rural sites. The
20 results of this gradient study were similar to those of a parallel OTC study. Also, the U.S.
21 forest service Forest Inventory and Analysis (FIA) program uses large-scale O3 exposure
22 patterns across the continental U.S. to study occurrences of foliar injury due to O3
23 exposure (Smith et al.. 2003) (Section 9.4.2). Finally, McLaughlin et al. (2007a; 2007b)
24 used spatial and temporal O3 gradients to study forest growth and water use in the
25 southern Appalachians. These studies found varying O3 exposures between years and
26 between sites.
9.2.6 Comparative Studies
27 All experimental approaches used to expose plants to O3 have strengths and weaknesses.
28 One potential weakness of laboratory, greenhouse, or field chamber studies is the
29 potential effect of the chamber on micrometeorology. In contrast, plume, FACE and
30 gradient systems are limited by the very small number of possible exposure levels
31 (almost always no more than two), small replication and an inability to reduce O3 levels
32 below ambient. In general, experiments that aim at characterizing the effect of a single
33 variable, e.g., exposure to O3, must not only manipulate the levels of that variable, but
34 also control potentially interacting variables and confounders, or else account for them.
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1 However, while increasing control of environmental variables makes it easier to discern
2 the effect of the variable of interest, it must be balanced with the ability to extend
3 conclusions to natural, non-experimental settings. More naturalistic exposure systems, on
4 the other hand, let interacting factors vary freely, resulting in greater unexplainable
5 variability. The various exposure methodologies used with O3 vary in the balance each
6 strikes between control of environmental inputs, closeness to the natural environment,
7 noisiness, and ability to make general inferences.
8 Studies have examined the comparability of results obtained though the various exposure
9 methodologies. As noted in the 1996 O3 AQCD, evidence from the comparative studies
10 of OTCs and from closed chamber and O3-exclusion exposure systems on the growth of
11 alfalfa (Medicago sativd) by Olszyk et al. (1986) suggested that, since significant
12 differences were found for fewer than 10% of the growth parameters measured, the
13 responses were, in general, essentially the same regardless of exposure system used, and
14 chamber effects did not significantly affect response. In 1988, Heagle et al. (1988)
15 concluded: "Although chamber effects on yield are common, there are no results showing
16 that this will result in a changed yield response to O3." A study of the effects of an
17 enclosure examined the responses of tolerant and sensitive white clover clones (Trifolium
18 repens) to ambient O3 in a greenhouse, open-top chamber, and ambient (no chamber)
19 plots (Heagle etal.. 1996). For individual harvests, greenhouse O3 exposure reduced the
20 forage weight of the sensitive clone 7 to 23% more than in OTCs. However, the response
21 in OTCs was the same as in ambient plots. Several studies have shown very similar
22 response of yield to O3 for plants grown in pots or in the ground, suggesting that even
23 such a significant change in environment does not alter the proportional response to O3,
24 providing that the plants are well watered (Heagle et al.. 1983; Heagle. 1979).
25 A few recent studies have compared results of O3 experiments between OTCs, FACE
26 experiments, and gradient studies. For example, a series of studies undertaken at Aspen
27 FACE (Isebrands et al.. 2001; Isebrands et al.. 2000) showed that O3 symptom
28 expression was generally similar in OTCs, FACE, and ambient O3 gradient sites, and
29 supported the previously observed variation among trembling aspen clones using OTCs
30 (Mankovska et al., 2005; Karnosky et al., 1999). In the SoyFACE experiment in Illinois,
31 soybean (Pioneer 93B15 cultivar) yield loss data from a two-year study was published
32 (Morgan et al.. 2006). This cultivar is a recent selection and, like most modern cultivars,
33 has been selected under an already high current O3 exposure. It was found to have
34 average sensitivity to O3 compared to 22 other cultivars tested at SoyFACE. In this
35 experiment, ambient hourly O3 concentrations were increased by approximately 20% and
36 measured yields were decreased by 15% in 2002 as a result of the increased O3 exposure
37 (Morgan et al.. 2006). To compare these results to chamber studies, Morgan et al. (2006)
3 8 calculated the expected yield loss from a linear relationship constructed from chamber
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1 data using seven-hour seasonal averages (Ashmore. 2002). They calculated an 8%
2 expected yield loss from the 2002 O3 exposure using that linear relationship. In another
3 study, Gregg et al. (2006. 2003) found similar O3 effects on cottonwood sapling biomass
4 growth along an ambient O3 gradient in the New York City area and a parallel OTC
5 study.
6 Finally, EPA conducted comparisons of exposure-response model predictions based on
7 OTC studies, and more recent FACE observations. These comparisons include yield of
8 annual crops, and biomass growth of trees. They are presented in section 9.6.3 of this
9 document.
9.3 Mechanisms Governing Vegetation Response to Ozone
9.3.1 Introduction
10 This section focuses on the effects of O3 stress on plants and their responses to that stress
11 on the molecular, biochemical and physiological levels. First, the pathway of O3 uptake
12 into the leaf and the initial chemical reactions occurring in the substomatal cavity and
13 apoplast will be described (Section 9.3.2); additionally, direct effects of O3 on the
14 stomatal apparatus will be discussed. Once O3 has entered the substomatal cavity and
15 apoplast, it is thought that the cell must be able to sense the presence of O3 or its
16 breakdown products in order to initiate the rapid changes in signaling pathways and gene
17 expression that have been measured in O3-treated plants. While it remains unclear exactly
18 how O3 and/or its breakdown products are sensed in the apoplast, much progress has been
19 made in examining several different mechanisms that may contribute both to sensing the
20 presence of O3 and its breakdown products, and also initiating a signal transduction
21 cascade, which will be described in Section 9.3.3.1. The next section focuses on changes
22 in gene and protein expression measured in plants exposed to O3, with particular
23 emphasis on results from transcriptome (all RNA molecules produced in a cell) and
24 proteome (all proteins produced in a cell) analyses (Section 9.3.3.2). Subsequently, the
25 role of phytohormones such as salicylic acid (SA), ethylene (ET), jasmonic acid (JA), and
26 abscisic acid (ABA) and their interactions in both signal transduction processes and in
27 determining plant response to O3 is discussed in Section 9.3.3.3. After O3 uptake and
28 sensing, some plants can respond to the oxidative stress with detoxification to minimize
29 damage. These mechanisms of detoxification, with particular emphasis on antioxidant
30 enzymes and metabolites, are reviewed in Section 9.3.4. The next section focuses on
31 changes in primary and secondary metabolism in plants exposed to O3, looking at
32 photosynthesis, respiration and several secondary metabolites, some of which may also
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1 act as antioxidants and protect the plant from oxidative stress (Section 9.3.5). For many
2 of these topics, information from the 2006 O3 AQCD has been summarized, as this
3 information is still valid and supported by more recent findings. For other topics, such as
4 genomics and proteomics, which have arisen due to the availability of new technologies,
5 the information is based solely on new publications with no reference to the 2006 O3
6 AQCD.
7 As Section 9.3 focuses on mechanisms underlying effects of O3 on plants and their
8 response to it, the conditions that are used to study these mechanisms do not always
9 reflect conditions that a plant may be exposed to in an agricultural setting or natural
10 ecosystem. The goal of many of these studies is to generate an O3 effect in a relatively
11 short period of time and not always to simulate ambient O3 exposures. Therefore, plants
12 are often exposed to unrealistically high O3 concentrations for several hours or days
13 (acute exposure), and only in some cases to ambient or slightly elevated O3
14 concentrations for longer time periods (chronic exposure). Additionally, the plant species
15 utilized in these studies are often not agriculturally important or commonly found as part
16 of natural ecosystems. Model organisms such as Arabidopsis thaliana are used frequently
17 as they are easy to work with, and mutants or transgenic plants are easy to develop or
18 have already been developed. Furthermore, the Arabidopsis genome has been sequenced,
19 and much is known about the molecular basis of many biochemical and cellular
20 processes.
21 Many of the studies described in this section focus on changes in the expression of genes
22 in O3-treated plants. Changes in gene expression (i.e., either up- or down-regulation of
23 gene expression) do not always translate into changes in protein quantity and/or activity,
24 as there are many levels of post-transcriptional and post-translational modifications
25 which impact protein quantity and activity. Many studies do not evaluate whether the
26 observed changes in gene expression lead to changes at the protein level and, therefore, it
27 is not always clear how relevant the changes in gene expression are in determining plant
28 response to O3. However, with the advent of proteomics, some very recent studies have
29 evaluated changes in protein expression for large numbers of proteins in O3 treated
30 plants, and the findings from these studies support the previous results regarding changes
31 in gene expression studies as a result of O3 exposure. The next step in the process is to
32 determine the implications of the measured changes occurring at the cellular level to
33 whole plants and ecosystems, which is an important topic of study which has not been
34 widely addressed.
35 The most significant new body of research since the 2006 O3 AQCD is on the
36 understanding of molecular mechanisms underlying how plants are affected by O3; a
37 significant number of recent studies reviewed here focus on changes in gene expression
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1 in plants exposed to elevated O3. Conclusions from the 2006 O3 AQCD have been
2 supported by these new studies, and the advent of new technologies has allowed for a
3 more comprehensive understanding of the mechanisms governing how plants are affected
4 byO3.
5 In summary, these new studies have increased knowledge of the molecular, biochemical
6 and cellular mechanisms occurring in plants in response to O3 by often using artificial
7 exposure conditions and model organisms. This information adds to the understanding of
8 the basic biology of how plants are affected by oxidative stress in the absence of any
9 other potential stressors. The results of these studies provide important insights, even
10 though they may not always directly translate into effects observed in other plants under
11 more realistic exposure conditions.
9.3.2 Ozone Uptake into the Leaf
12 Appendix AX9.2.3 of the 2006 O3 AQCD clearly described the process by which O3
13 enters plant leaves through open stomata (U.S. EPA. 2006b). This information continues
14 to be valid and is only summarized here.
15 Stomata provide the principal pathway for O3 to enter and affect plants (Massman and
16 Grantz. 1995; Fuentes et al., 1992; Reich. 1987; Leuning et al., 1979). Ozone moves into
17 the leaf interior by diffusing through open stomata, and environmental conditions which
18 promote high rates of gas exchange will favor the uptake of the pollutant by the leaf.
19 Factors that may limit uptake include boundary layer resistance and the size of the
20 stomatal aperture (Figure 9-2) (U.S. EPA. 2006b). Once inside the substomatal cavity, O3
21 is thought to rapidly react with the aqueous apoplast to form breakdown products known
22 as reactive oxygen species (ROS), such as hydrogen peroxide (H2O2), superoxide (O2 ),
23 hydroxyl radicals (HO) and peroxy radicals (HO2) (Figure 9-3). Hydrogen peroxide is
24 not only a toxic breakdown product of O3, but has been shown to function as a signaling
25 molecule, which is activated in response to both biotic and abiotic stressors. The role of
26 H2O2 in signaling was described in detail in the 2006 O3 AQCD. Additional organic
27 molecules present in the apoplast or cell wall, such as those containing double bonds or
28 sulfhydryls that are sensitive to oxidation, could also be converted to oxygenated
29 molecules after interacting with O3 (Figure 9-4). These reactions are not only pH
30 dependent, but are also influenced by the presence of other molecules in the apoplast
31 (U.S. EPA. 2006b). The 2006 O3 AQCD provided a comprehensive summary of what is
32 known about the possible interactions of O3 with other biomolecules (U.S. EPA. 2006b).
33 It is in the apoplast that initial detoxification reactions by antioxidant metabolites and
34 enzymes take place, and these initial reactions are critical to reduce concentrations of the
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1 oxidative breakdown products of O3; these reactions are described in more detail in
2 Section 9.3.4 of this document.
9.3.2.1 Changes in Stomatal Function
3 The effects of O3 exposure on stomatal conductance have been reviewed in detail in
4 previous O3 AQCDs. Although the nature of these effects depends upon many different
5 factors, including the plant species, concentration and duration of the O3 exposure, and
6 prevailing meteorological conditions, stomatal conductance is often negatively affected
7 by plant exposure to O3 CWittig etal.. 2007). Decreases in conductance have been shown
8 to result from declines in photosynthetic carboxylation capacity, leading to a buildup of
9 CO2 in the substomatal cavity and subsequent stomatal closure (Wittig et al.. 2007).
10 However, results from the use of Arabidopsis mutants and new technologies, which allow
11 for analysis of guard cell function in whole plants rather than in isolated guard cells or
12 epidermal peels, suggest that O3 may also have a direct impact on stomatal guard cells,
13 leading to alterations in stomatal conductance. The use of a new simultaneous O3
14 exposure/gas exchange device has demonstrated that exposure of Arabidopsis ecotypes
15 Col-0 and Ler to 150 ppb O3 resulted in a 60-70% decline in stomatal conductance within
16 9-12 minutes of beginning the exposure. Twenty to thirty minutes later, stomatal
17 conductance had returned to its initial value, even with continuing exposure to O3,
18 indicating a rapid direct effect of O3 on stomatal function (Kollist et al., 2007). This
19 transient decrease in stomatal conductance was not observed in the abscisic acid
20 insensitive (ABI2) Arabidopsis mutant. As the ABE protein is thought to regulate the
21 signal transduction process involved in stomatal response downstream of ROS
22 production, the authors suggest that the transient decrease in stomatal conductance in the
23 Col-0 and Ler ecotypes results from the biological action of ROS in transducing signals,
24 rather than direct physical damage to guard cells by ROS (Kollist et al.. 2007). This rapid
25 transient decrease in stomatal conductance was also not observed when exposing the
26 Arabidopsis mutant slacl (slow anion channel-associated 1) to 200 ppb O3 (Vahisalu et
27 al.. 2008). The SLAC1 protein was shown to be essential for guard cell slow anion
28 channel functioning and for stomatal closure in response to O3. Based on additional
29 studies using a variety of Arabidopsis mutants impaired in various aspects of stomatal
30 function, Vahisalu et al. (2010) suggest that the presence of ROS in the guard cell
31 apoplast (formed either by O3 breakdown or through ROS production from NADPH
32 oxidase activity) leads to the activation of a signaling pathway in the guard cells, which
33 includes SLAC1, and results in stomatal closure.
34 A review by McAinsh et al. (2002) discusses the role of calcium as a part of the signal
35 transduction pathway involved in regulating stomatal responses to pollutant stress. A
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1 number of studies in this review provide some evidence that exposure to O3 increases the
2 cytosolic free calcium concentration ([Ca2+]cyt) in guard cells, which may result in an
3 inhibition of the plasma membrane inward-rectifying K+ channels in guard cells, which
4 allow for the K+ uptake needed for stomatal opening (McAinsh et al.. 2002; Torsethaugen
5 et al., 1999). This would compromise the ability of the stomata to respond to various
6 stimuli, including light, CO2 concentration and drought. Pei et al. (2000) reported that the
7 presence of H2O2 activated Ca2+ -permeable channels, which mediate increases in
8 [Ca2+]cyt in guard cell plasma membranes of Arabidopsis. They also determined that
9 abscisic acid (ABA) induced H2O2 production in guard cells, leading to ABA-induced
10 stomatal closure via activation of the membrane Ca2+ channels. Therefore, it is possible
11 that H2O2, a byproduct of O3 breakdown in the apoplast, could disrupt the Ca2+-ABA
12 signaling pathway that is involved in regulating stomatal responses (McAinsh et al..
13 2002). The studies described here provide some evidence to suggest that O3 and its
14 breakdown products can directly affect stomatal functioning by impacting the signal
15 transduction pathways which regulate guard cells. Stomatal sluggishness has been
16 described as a delay in stomatal response to changing environmental conditions in
17 sensitive species exposed to higher concentrations and/or longer-term O3 exposures
18 (Paoletti and Grulke. 2010. 2005; McAinsh et al.. 2002). It is possible that the signaling
19 pathways described above could be involved in mediating this stomatal sluggishness in
20 some plant species under certain O3 exposure conditions (Paoletti and Grulke. 2005;
21 McAinsh et al.. 2002).
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Light
Cuticle
Epidermis
Pallisade
Mesophyll
Spongy
Mesophyll
Epidermis
Cuticle
1?
mrnrn
Vascular
System
C0=[C02]-
Figure 9-2 The microarchitecture of a dicot leaf. While details among species
vary, the general overview remains the same. Light that drives
photosynthesis generally falls upon the upper (adaxial) leaf
surface. Carbon dioxide and ozone enter through the stomata on
the lower (abaxial) leaf surface, while water vapor exits through the
stomata (transpiration).
a.
Stiperoxide
Hydroxyt
Radical
b.
H2°2
HO- H2O2
Pe/oxy^
Radical
Figure 9-3 Possible reactions of ozone within water, (a) Ozone reacts at the
double bonds to form carbonyl groups, (b) Under certain
circumstances, peroxides are generated.
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9-15
September 2011
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a.
2.
O
OH-
Crigee
Mechanism
H2C = CH2
H2C = CH2
0 O
\ I
OH
\
H
HC=O
0
ii
HC-OH
NO
H2C=CH2
ONO2
b.
Source: Adapted from Mudd (1996).
CH(OH)CH O2H
CH(OH)CH 02H
OH
\
0=C CH(OH)CH02H
CHO , CHO
Further Oxidation
Figure 9-4 The Crigee mechanism of ozone attack of a double bond, (a) The
typical Crigee mechanism is shown in which several reactions
paths from the initial product is shown, (b) Typical reaction of
ascorbic acid with ozone.
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9.3.3 Cellular to Systemic Responses
9.3.3.1 Ozone Sensing and Signal Transduction
1 New technologies allowing for large-scale analysis of oxidative stress-induced changes in
2 gene expression have facilitated the study of signal transduction processes associated
3 with the perception and integration of responses to the stress. Many of these studies have
4 been conducted using Arabidopsis or tobacco plants, for which a variety of mutants are
5 available and/or which can be easily genetically modified to generate either loss-of-
6 function or over-expressing genotypes. Several comprehensive review articles provide an
7 overview of what is known of O3-induced signal transduction processes and how they
8 may help to explain differential sensitivity of plants to the pollutant (Ludwikow and
9 Sadowski. 2008; Baier et al.. 2005; Kangasjarvi et al.. 2005). Additionally, analysis of
10 several studies of transcriptome changes has also allowed for the compilation of these
11 data to determine an initial time-course for O3-induced activation of various signaling
12 compounds (Kangasjarvi et al.. 2005).
13 A number of different mechanisms for plant sensing of O3 have been proposed; however,
14 there is still much that is not known about this process. Some of the earliest events that
15 occur in plants exposed to O3 have been described in the guard cells of stomata. Reactive
16 oxygen species were observed in the chloroplasts of guard cells in the O3 tolerant Col-0
17 Arabidopsis thaliana ecotype plants within 5 minutes of plant exposure to 350 ppb O3
18 (Joo et al.. 2005). Reactive oxygen species from the breakdown of O3 in the apoplast are
19 believed to activate GTPases (G-proteins), which, in turn, activate several intracellular
20 sources of ROS, including ROS derived from the chloroplasts. G-proteins are also
21 believed to play a role in activating membrane-bound NADPH oxidases to produce ROS
22 and, as a result, propagate the oxidative burst to neighboring cells (Joo et al.. 2005).
23 Therefore, G-proteins are recognized as important molecules involved in plant responses
24 to O3 and may play a role in perceiving ROS from the breakdown of O3 in the apoplast
25 (Kangasjarvi et al.. 2005; Booker etal., 2004b).
26 A change in the redox state of the plant and the oxidation of sensitive molecules in itself
27 may represent a means of perception and signaling of oxidative stress in plants.
28 Disulfide-thiol conversions in proteins and the redox state of the glutathione pool may be
29 important components of redox sensing and signal transduction (Foyer and Noctor.
30 2005a. b).
31 Calcium (Ca2+) has also been implicated in the transduction of signals to the nucleus in
32 response to oxidative stress. The influx of Ca2+ from the apoplast into the cell occurs
33 early during plant exposure to O3, and it is thought to play a role in regulating the activity
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1 of protein kinases, which are discussed below (Baier et al.. 2005; Hamel et al.. 2005).
2 Calcium channel blockers inhibited O3-induced activation of protein kinases in tobacco
3 suspension cells exposed to 500 ppb O3 for 10 minutes, indicating that the opening of
4 Ca2+ channels is an important upstream signaling event or that the as yet unknown
5 upstream process has a requirement for Ca2+ (Samuel et al.. 2000).
6 Further transmission of information regarding the presence of ROS to the nucleus t
7 involves mitogen-activated protein kinases (MAPK), which phosphorylate proteins and
8 activate various cellular responses (Hamel et al.. 2005). Mitogen-activated protein
9 kinases are induced in several different plant species in response to O3 exposure,
10 including tobacco (Samuel et al.. 2005), Arabidopsis (Ludwikow et al.. 2004), the shrub
11 Phillyrea latifolia (Paolacci et al.. 2007) and poplar (Hamel et al.. 2005). Disruption of
12 these signal transduction pathways by over-expressing or suppressing MAP kinase
13 activity in different Arabidopsis and tobacco lines resulted in increased plant sensitivity
14 to O3 (Miles et al., 2005; Samuel and Ellis, 2002). Additionally, greater O3 tolerance of
15 several Arabidopsis ecotypes was correlated with greater up-regulation of MAP kinase
16 signaling pathways upon O3 exposure than in more sensitive Arabidopsis ecotypes (Li et
17 al.. 2006b; Mahalingam et al.. 2006; Overmyer et al.. 2005). indicating that determination
18 of plant sensitivity and plant response to O3 may, in part, be determined not only by
19 whether these pathways are turned on, but also by the magnitude of the signals moving
20 through these communication channels.
21 In conclusion, experimental evidence suggests that there are likely several different
22 mechanisms by which the plant senses the presence of O3 or its breakdown products.
23 These mechanisms may vary by species or developmental stage of the plant, or may co-
24 exist and be activated by different exposure conditions. Calcium and protein kinases are
25 likely involved in relaying information about the presence of the stressor to the nucleus
26 and other cellular compartments as a first step in determining whether and how the plant
27 will respond to the stress.
9.3.3.2 Gene and Protein Expression Changes in Response to
Ozone
28 The advent of DNA microarray technology has allowed for the study of gene expression
29 in cells on a large scale. Rather than assessing changes in gene expression of individual
30 genes, DNA microarrays facilitate the evaluation of entire transcriptomes, providing a
31 comprehensive picture of simultaneous alterations in gene expression. In addition, these
32 studies have provided more insight into the complex interactions between molecules, how
33 those interactions lead to the communication of information in the cell (or between
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1 neighboring cells), and which role these interactions play in determining tolerance or
2 sensitivity and how a plant may respond to stresses such as O3 (Ludwikow and
3 Sadowski. 2008). Transcriptome analysis of O3-treated plants has been performed in
4 several species, including Arabidopsis thaliana (Li et al.. 2006b: Tosti et al.. 2006;
5 Heidenreich et al., 2005; Mahalingam et al., 2005; Tamaoki et al., 2003). pepper
6 (Capsicum annuum) (Lee and Yun. 2006). clover (Medicago truncatuld) (Puckette et al..
7 2008). Phillyrea latifolla (Paolacci et al.. 2007). poplar (Street et al.. 2011). and European
8 beech (Fagus sylvatica) (Olbrich etal.. 2010: Olbrich et al.. 2009; Olbrich et al.. 2005).
9 In some cases, researchers compared transcriptomes of two or more cultivars, ecotypes or
10 mutants that differed in their sensitivity to O3 (Puckette et al.. 2008; Rizzo et al.. 2007;
11 Lee and Yun. 2006; Li et al.. 2006b; Tamaoki et al.. 2003). Species, O3 exposure
12 conditions (concentration, duration of exposure) and sampling times varied significantly
13 in these studies. However, functional classification of the genes that were either up- or
14 down-regulated by plant exposure to O3 exhibited common trends. Genes involved in
15 plant defense, signaling and those associated with the synthesis of plant hormones and
16 secondary metabolism were generally up-regulated, while those related to photosynthesis
17 and general metabolism were typically down-regulated in O3-treated plants (Puckette et
18 al.. 2008; Lee and Yun. 2006; Li et al.. 2006b; Tosti et al.. 2006; Olbrich et al.. 2005;
19 Tamaoki etal.. 2003).
20 Analysis of the transcriptome has been used to evaluate differences in gene expression
21 between O3 sensitive and tolerant plants. In pepper, 67% of the 180 genes studied that
22 were affected by O3 were differentially regulated in the sensitive and tolerant cultivars.
23 At both 0 hours and 48 hours after a 3-day exposure at 150 ppb, O3 responsive genes
24 were either up- or down-regulated more markedly in the sensitive than in the tolerant
25 cultivar (Lee and Yun. 2006). Transcriptome analysis also revealed differences in timing
26 and magnitude of changes in gene expression between sensitive and tolerant clovers.
27 Acute exposure (300 ppb O3 for 6 hours) led to the production of an oxidative burst in
28 both clovers (Puckette et al.. 2008). However, the sensitive Jemalong cultivar exhibited a
29 sustained ROS burst and a concomitant down-regulation of defense response genes at
30 12 hours after the onset of exposure, while the tolerant JE 154 accession showed much
31 more rapid and large-scale transcriptome changes than the Jemalong cultivar (Puckette et
32 al.. 2008).
33 Arabidopsis ecotypes WS and Col-0 were exposed to 1.2 x ambient O3 concentrations for
34 8-12 days at the Soy FACE site (Li et al.. 2006b). The sensitive WS ecotype showed a far
35 greater number of changes in gene expression in response to this low-level O3 exposure
36 than the tolerant Col-0 ecotype. In a different study, exposure of the WS ecotype to
37 300 ppb O3 for 6 hours showed a rapid induction of genes leading to cell death, such as
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1 proteases, and down-regulation or inactivation of cell signaling genes, demonstrating an
2 ineffective defense response in this O3 sensitive ecotype (Mahalingam et al.. 2006).
3 The temporal response of plants to O3 exposure was evaluated in the Arabidopsis Col-0
4 ecotype during a 6-h exposure at 350 ppb O3 and for 6 hours after the exposure was
5 completed. Results of this study, shown in Figure 9-5, indicate that genes associated with
6 signal transduction and regulation of transcription were in the class of early up-regulated
7 genes, while genes associated with redox homeostasis and defense/stress response were
8 in the class of late up-regulated genes (Mahalingam et al.. 2005).
9 A few studies have been conducted to evaluate transcriptome changes in response to
10 longer term chronic O3 exposures in woody plant species. Longer term exposures
11 resulted in the up-regulation of genes associated with secondary metabolites, including
12 isoprenoids, polyamines and phenylpropanoids in 2-year-old seedlings of the
13 Mediterranean shrub Phillyrea latifolia exposed to 110 ppb O3 for 90 days (Paolacci et
14 al.. 2007). In 3-year-old European beech saplings exposed to O3 for 20 months, with
15 monthly average twice ambient O3 concentrations ranging from 11 to 80 ppb, O3-induced
16 changes in gene transcription were similar to those observed for herbaceous species
17 (Olbrich et al.. 2009). Genes encoding proteins associated with plant stress response,
18 including ethylene biosynthesis, pathogenesis-related proteins and enzymes detoxifying
19 ROS, were up-regulated. Some genes associated with primary metabolism, cell structure,
20 cell division and cell growth were reduced (Olbrich et al.. 2009). In a similar study using
21 adult European beech trees, it was determined that the magnitude of the transcriptional
22 changes described above was far greater in the saplings than in the adult trees exposed to
23 the same O3 concentrations for the same time period (Olbrich etal.. 2010).
24 The results from transcriptome studies described above have been substantiated by results
25 from proteome analysis in rice, poplar, European beech, wheat, and soybean. Exposure of
26 soybean to 120 ppb O3 for 12-h/day for 3 days in growth chambers resulted in decreases
27 in the quantity of proteins associated with photosynthesis, while proteins involved with
28 antioxidant defense and carbon metabolism increased (Ahsan etal.. 2010). Young poplar
29 plants exposed to 120 ppb O3 in a growth chamber for 35 days also showed significant
30 changes in proteins involved in carbon metabolism (Bohler et al.. 2007). Declines in
31 enzymes associated with carbon fixation, the Calvin cycle and photosystem II were
32 measured, while ascorbate peroxidase and enzymes associated with glucose catabolism
33 increased in abundance. In another study to determine the impacts of O3 on both
34 developing and fully expanded poplar leaves, young poplars were exposed to 120 ppb O3
35 for 13-h per day for up to 28 days (Bohler et al.. 2010). Impacts on protein quantity only
36 occurred after the plants had been exposed to O3 for 14 days, and at this point in time,
37 several Calvin cycle enzymes were reduced in quantity, while the effects on the light
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1 reactions appeared later, at 21 days after beginning treatment. Some of the antioxidant
2 enzymes increased in abundance with O3 treatment, while others (ascorbate peroxidase)
3 did not. In relationship to leaf expansion, it was shown that O3 did not affect protein
4 quantity until leaves had reached full expansion, after about 7 days (Bohler et al.. 2010).
5 Two-week-old rice seedlings exposed to varying levels of O3 (4, 40, 80, 120 ppb) in a
6 growth chamber for 9 days showed reductions in quantities of proteins associated with
7 photosynthesis and energy metabolism, and increases in some antioxidant and defense
8 related proteins (Fenget al., 2008a). A subsequent study of O3-treated rice seedlings
9 (exposed to 200 ppb O3 for 24-h) focusing on the integration of transcriptomics and
10 proteomics, supported and further enhanced these results (Cho et al., 2008). The authors
11 found that of the 22,000 genes analyzed from the rice genome, 1,535 were differentially
12 regulated by O3. Those differentially regulated genes were functionally categorized as
13 transcription factors, MAPK cascades, those encoding for enzymes involved in the
14 synthesis of jasmonic acid (JA),ethylene (ET), shikimate, tryptophan and lignin, and
15 those involved in glycolysis, the citric acid cycle, oxidative respiration and
16 photosynthesis. The authors determined that the proteome and metabolome (all small
17 molecule metabolites in a cell) analysis supported the results of the transcriptome
18 changes described above (Cho et al.. 2008). This type of study, which ties together results
19 from changes in gene expression, protein quantity and activity, and metabolite levels,
20 provides the most complete picture of the molecular and biochemical changes occurring
21 in plants exposed to a stressor such as O3.
22 Sarkar et al. (2010) compared proteome s of two cultivars of wheat grown in OTCs at
23 several O3 concentrations, including filtered air, ambient O3 (mean concentration
24 47 ppb), ambient +10 ppb and ambient + 20 ppb for 5-h/day for 50 days. Declines in the
25 rate of photosynthesis and stomatal conductance were related to decreases in proteins
26 involved in carbon fixation and electron transport and increased proteolysis of
27 photosynthetic proteins such as the large subunit of ribulose-l,6-bisphosphate
28 carboxylase/oxygenase (Rubisco). Enzymes that take part in energy metabolism, such as
29 ATP synthesis, were also down-regulated, while defense/stress related proteins were up-
30 regulated in O3 treated plants. In comparing the two wheat cultivars, Sarkar et al. (2010)
31 found that while the qualitative changes in protein expression between the two cultivars
32 were similar, the magnitude of these changes differed between the sensitive and tolerant
33 wheat cultivars. Greater foliar injury and a smaller decline in stomatal conductance was
34 observed in the sensitive cultivar as compared to the more tolerant cultivar, along with
3 5 greater losses in photosynthetic enzymes and higher quantities of antioxidant enzymes.
36 Results from a three year exposure of European beech saplings to elevated O3 (AOT 40
37 value was 52.6 ul 1-1-h for 2006 when trees were sampled) supported the results from the
38 short-term exposure studies described above (Kerner et al., 2011). The O3 treatment of
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1 the saplings resulted in reductions in enzymes associated with the Calvin cycle, which
2 could lead to reduced carbon fixation. Enzymes associated with carbon
3 metabolism/catabolism were increased, and quantities of starch and sucrose were reduced
4 in response to the O3 treatment in these trees, indicating a potential impact of O3 on
5 overall carbon metabolism in long-term exposure conditions (Kerner et al., 2011).
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(A)
Signaling
Transcription
Redox homeostasis
Dcfcnsc/slrcss response
PR proteins
t
12 hr
12 hr
Photosynthesis
Source: Used with permission from Springer (Mahalingam et al.. 2005).
(A) Temporal profile of the oxidative stress response to ozone. The biphasic ozone-induced oxidative burst is represented in black,
with the ROS control measurements shown as a broken line. Average transcript profiles are shown for early up-regulated genes
(yellow, peaks at 0.5-1 hours), and the 3 hours (blue), 4.5 hours (red) and 9-12 hours (green) late up-regulated genes and for the
down-regulated genes coding for photosynthesis proteins (brown). (B) Diagrammatic representation of redox regulation of the
oxidative stress response.
Figure 9-5 Composite diagram of major themes in the temporal evolution of
1 the genetic response to OZOne Stress. All of these studies describe common
2 trends for changes in gene and protein expression which occur in a variety of plant
3 species exposed to O3. While genes associated with carbon assimilation and general
4 metabolism are typically down-regulated, genes associated with signaling, catabolism,
5 and defense are up-regulated. The magnitude of these changes in gene and protein
6 expression appears to be related to plant species, age and their sensitivity or tolerance to
7 03.
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9.3.3.3 Role of Phytohormones in Plant Response to Ozone
1 Many studies of O3 effects on plants have analyzed the importance of plant hormones
2 such as SA, ET and JA in determining plant response to O3; some of the roles of these
3 hormones were described in the 2006 O3 AQCD. Transcriptome analysis and the use of a
4 variety of mutants have allowed for further elucidation of the complex interactions
5 between SA, ET, JA and the role of abscisic acid (ABA) in mediating plant response to
6 O3 (Ludwikow and Sadowski. 2008). In addition to their roles in signaling pathways,
7 phytohormones also appear to regulate, and be regulated by, the MAPK signaling
8 cascades described previously. Most evidence suggests that while ET and SA are needed
9 to develop O3-induced leaf lesions, JA acts antagonistically to SA and ET to limit the
10 lesions (Figure 9-6) (Kangasjarvi et al.. 2005).
11 The rapid production of ET in O3 treated plants has been described in many plant species
12 and has been further characterized through the use of a variety of mutants that either
13 over-produce or are insensitive to ET. Production of stress ET in O3-treated plants, which
14 is thought to be part of a wounding response, was found to be correlated to the degree of
15 injury development in leaves (U.S. EPA. 2006b). More recent studies have supported
16 these conclusions and have also focused on the interactions occurring between several
17 oxidative-stress induced phytohormones. Yoshida et al. (2009) determined that ET likely
18 amplifies the oxidative signal generated by ROS, thereby promoting lesion formation. By
19 analyzing the O3-induced transcriptome of several Arabidopsis mutants of the Col-0
20 ecotype, Tamaoki et al. (2003) determined that at 12 hours after initiating the O3
21 exposure (200 ppb for 12 hours), the ET and JA signaling pathways were the main
22 pathways used to activate plant defense responses, with a lesser role for SA. The authors
23 also demonstrated that low levels of ET production could stimulate the expression of
24 defense genes, rather than promoting cell death which occurs when ET production is
25 high. Tosti et al. (2006) supported these findings by showing that plant exposure to O3
26 not only results in activation of the biosynthetic pathways of ET, JA and SA, but also
27 increases the expression of genes related to the signal transduction pathways of these
28 phytohormones in O3-treated Arabidopsis plants (300 ppb O3 for 6 hours). Conversely, in
29 the O3 sensitive Ws ecotype, its sensitivity may, in part, be due to intrinsically high ET
30 levels leading to SA accumulation, and the high ET and SA may act to repress JA-
31 associated genes, which would serve to inhibit the spread of lesions (Mahalingam et al..
32 2006). Ogawa et al. (2005) found that increases in SA in O3-treated plants leads to the
33 formation of leaf lesions in tobacco plants exposed to 200 ppb O3 for 6 hours.
34 Furthermore, in transgenic tobacco plants with reduced levels of ET production in
35 response to O3 exposure, several genes encoding for enzymes in the biosynthetic pathway
36 of SA were suppressed, suggesting that SA levels are, in part, controlled by ET in the
37 presence of O3.
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1 Exposure of the Arabidopsis mutant rcdl to acute doses of O3 (250 ppb O3 for 8-h/day
2 for 3 days) resulted in programmed cell death (PCD) and the formation of leaf lesions
3 (Overmyer et al., 2000). They determined that the observed induction of ET synthesis
4 promotes cell death, and that ET perception and signaling are required for the
5 accumulation of superoxide, which leads to cell death and propagation of lesions. .
6 Jasmonic acid, conversely, contains the spread of leaf lesions (Overmyer et al.. 2000).
7 Transcriptome analysis of several Arabidopsis mutants, which are insensitive to SA, ET
8 and JA, exposed to 12-h of 200 ppb O3 showed that approximately 78 of the up-regulated
9 genes measured in this study were controlled by ET and JA signaling pathways, while SA
10 signaling pathways were suggested to antagonize ET and JA pathways (Tamaoki et al..
11 2003). In a subsequent transcriptome study on the Col-0 ecotype exposed to 150 ppb O3
12 for 48-h, JA and ET synthesis was down-regulated, while SA was up-regulated in O3-
13 treated plants. In cotton plants exposed to a range of O3 concentrations (0-120 ppb) and
14 methyl jasmonate (MeJA), Grantz et al. (2010a) determined that exogenous applications
15 of MeJA did not protect plants from chronic O 3 exposure.
16 Abscisic acid has been investigated for its role in regulating stomatal aperture and also
17 for its contribution to signaling pathways in the plant. The role of ABA and the
18 interaction between ABA and H2O2 in O3-induced stomatal closure was described in the
19 2006 O3 AQCD. More recently, it was determined that synthesis of ABA was induced in
20 O3-treated Arabidopsis plants (250-350 ppb O3 for 6 hours), with a more pronounced
21 induction in the O3 sensitive rcd3 mutant as compared to the wildtype Col-0 (Overmyer
22 et al.. 2008). The rcd3 mutant also exhibited a lack of O3-induced stomatal closure, and
23 the RCD3 protein has been shown to be required for slow anion channels (Overmyer et
24 al.. 2008) (see Section 9.3.4.1). Ludwikow et al. (2009) used Arabidopsis ABIltd
25 mutants, in which a key negative regulator of ABA action (abscisic acid insensitive 1
26 protein phosphatase 2C) has been knocked out, to examine O3 responsive genes in this
27 mutant compared to the Arabidopsis Col-0. Results of this study indicate a role for ABU
28 in negatively regulating the synthesis of both ABA and ET in O3-treated plants (350 ppb
29 O3 for 9 hours). Additionally, ABU may stimulate JA-related gene expression, providing
30 evidence for an antagonistic interaction between ABA and JA signaling pathways
31 (Ludwikow etal.. 2009).
32 Nitric oxide (NO) has also been shown to play a role in regulating gene expression in
33 plants in response to O3 exposure. However, little is known to date about NO and its role
34 in the complex interactions of molecules in response to O3. Exposure of tobacco to O3
35 (150 ppb for 5 hours) stimulated NO and NO-dependent ET production, while NO
36 production itself did not depend on the presence of ET (Ederli etal.. 2006). Analysis of
37 O3-treated Arabidopsis indicated the possibility of a dual role for NO in the initiation of
38 cell death and later lesion containment (Ahlfors et al., 2009).
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1
2
3
4
5
6
7
While much work remains to be done to better elucidate how plants sense O3, what
determines their sensitivity to the pollutant and how they might respond to it, it is clear
that the mechanism for O3 sensing and signal transduction is very complex. Many of the
phytohormones and other signaling molecules thought to be involved in these processes
are interactive and depend upon a variety of other factors, which could be either internal
or external to the plant. This results in a highly dynamic and complex system, capable of
resulting in a spectrum of plant sensitivity to oxidative stress and generating a variety of
plant responses to that stress.
ozone
Cell
death
Source: Used with permission from Blackwell Publishing Ltd. (Kangasiarvi et al.. 2005).
Ozone-derived radicals induce endogenous ROS production (1) which results in salicylic acid (SA) accumulation and programmed
cell death; (2) Cell death triggers ethylene (ET) production, which is required for the continuing ROS production responsible for the
propagation of cell death; (3) Jasmonates counteract the progression of the cycle by antagonizing the cell death promoting function
of SA and ET; (4) Abscisic acid (ABA) antagonizes ET function in many situations and might also have this role in ozone-induced
cell death; (5) Mutually antagonistic interactions between ET, SAand jasmonic acid (JA) are indicated with red bars.
Figure 9-6 The oxidative cell death cycle. Detoxification
9
10
11
12
13
14
15
9.3.4.1 Overview of Ozone-Induced Defense Mechanisms
Plants are exposed to an oxidizing environment on a continual basis, and many reactions
that are part of the basic metabolic processes, such as photosynthesis and respiration,
generate ROS. As a result, there is an extensive and complex mechanism in place to
detoxify these oxidizing radicals, including both enzymes and metabolites, which are
located in several locations in the cell and also in the apoplast of the cell. As O3 enters
the leaf through open stomata, the first point of contact of O3 with the plant is likely in
the apoplast, where it breaks down to form oxidizing radicals such as H2O2, O2, HO- and
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September 2011
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1 HO 2. Another source of oxidizing radicals is an oxidative burst, generated by a
2 membrane-bound NADPH oxidase enzyme, which is recognized as an integral
3 component of the plant's defense system against pathogens (Schraudner et al., 1998).
4 Antioxidant metabolites and enzymes located in the apoplast are thought to form a first
5 line of defense by detoxifying O3 and/or the ROS that are formed as breakdown products
6 of O3 (Section 9.3.2.). However, even with the presence of several antioxidants,
7 including ascorbate, the redox buffering capacity of the apoplast is far less than that of
8 the cytoplasm, as it lacks the regeneration systems necessary to retain a reduced pool of
9 antioxidants (Foyer and Noctor. 2005b).
10 Redox homeostasis is regulated by the presence of a pool of antioxidants, which are
11 typically found in a reduced state and detoxify ROS produced by oxidases or electron
12 transport components. As ROS increase due to environmental stress such as O3, it is
13 unclear whether the antioxidant pool can maintain its reduced state (Foyer and Noctor.
14 2005b). As such, not only the quantity and types of antioxidant enzymes and metabolites
15 present, but also the cellular ability to regenerate those antioxidants are important
16 considerations in mechanisms of plant tolerance to oxidative stress (Dizengremel et al.,
17 2008). Molecules such as glutathione (GSH), thioredoxins and NADPH play very
18 important roles in this regeneration process; additionally, it has been hypothesized that
19 alterations in carbon metabolism would be necessary to supply the needed reducing
20 power for antioxidant regeneration (Dizengremel et al.. 2008).
9.3.4.2 Role of Antioxidants in Plant Defense Responses
21 Ascorbate has been the focus of many different studies as an antioxidant metabolite that
22 protects plants from exposure to O3. It is found in several cellular locations, including the
23 chloroplast, the cytosol and the apoplast (Noctor and Foyer. 1998). Ascorbate is
24 synthesized in the cell and transported to the apoplast. Apoplastic ascorbate can be
25 oxidized to dehydroascorbate (DHA) with exposure to O3 and is then transported back to
26 the cytoplasm. Here, DHA is reduced to ascorbate by the enzyme dehydroascorbate
27 reductase (DHAR) and reduced GSH, which is part of the ascorbate-glutathione cycle
28 (Noctor and Foyer. 1998). Many studies have focused on evaluating whether ascorbate is
29 the primary determining factor in differential sensitivity of plants to O3. An evaluation of
30 several species of wildflowers in Great Smoky Mountains National Park showed a
31 correlation between higher quantities of reduced apoplastic ascorbate and lower levels of
32 foliar injury from O3 exposure in the field in tall milkweed plants (Asclepsias exaltata L.)
33 (Burkey et al.. 2006; Souza et al.. 2006). Cheng et al. (2007) exposed two soybean
34 cultivars to elevated O3 (77 ppb) and filtered air for 7-h/day for 6 days. The differences in
35 sensitivity between the two cultivars could not be explained by differential O3 uptake or
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1 by the fraction of reduced ascorbate present in the apoplast. However, total antioxidant
2 capacity of the apoplast was twofold higher in the tolerant Essex cultivar as compared to
3 the sensitive Forrest cultivar, indicating that there may be other compounds in the leaf
4 apoplast that scavenge ROS. D'Haese et al. (2005) exposed the NC-S (sensitive) and NC-
5 R (resistant) clones of white clover (Trifolium repens) to 60 ppb O3 for 7-h/day for
6 5 days in environmental chambers. Surprisingly, the NC-S clone had a higher constitutive
7 concentration of apoplastic ascorbate with a higher redox status than the NC-R clone.
8 However, the redox status of symplastic GSH was higher in NC-R, even though the
9 concentration of GSH was not higher than in NC-S. In addition, total symplastic
10 antioxidative capacity was not a determining factor in differential sensitivity between
11 these two clones. Severino et al. (2007) also examined the role of antioxidants in the
12 differential sensitivity of the two white clover clones by growing them in the field for a
13 growing season and then exposing them to elevated O3 (100 ppb for 8-h/day for 10 days)
14 in OTC at the end of the field season. The NC-R clone had greater quantities of total
15 ascorbate and total antioxidants than the NC-S clone at the end of the experiment. In snap
16 bean, plants of the O3 tolerant Provider cultivar had greater total ascorbate and more
17 ascorbate in the apoplast than the sensitive SI56 cultivar after exposure to 71 ppb O3 for
18 10 days in OTC (Burkey et al.. 2003). While most of the apoplastic ascorbate was in the
19 oxidized form, the ratio of reduced ascorbate to total ascorbate was higher in Provider
20 than S156, indicating that Provider is better able to maintain this ratio to maximize plant
21 protection from oxidative stress. Exposure of two wheat varieties to ambient (7-h average
22 44 ppb O3) and elevated (7-h average 56 ppb O3) for 60 days in open-air field conditions
23 showed higher concentrations of reduced ascorbate in the apoplast in the tolerant Y16
24 variety than the more sensitive Y2 variety, however no varietal differences were seen in
25 the decrease in reduced ascorbate quantity in response to O3 exposure (Feng etal.. 2010).
26 There is much evidence that supports an important role for ascorbate, particularly
27 apoplastic ascorbate, in protecting plants from oxidative stressors such as O3; however, it
28 is also clear that there is much variation in the importance of ascorbate for different plant
29 species and differing exposure conditions. Additionally, the work of several authors
30 suggests that there may be other compounds in the apoplast which have the capacity to
31 act as antioxidants.
32 While the quantities of antioxidant metabolites such as ascorbate are an important
33 indicator of plant tolerance to O3, the ability of the plant to recycle oxidized ascorbate
34 efficiently also plays a large role in determining the plant's ability to effectively protect
35 itself from sustained exposure to oxidative stress. Tobacco plants over-expressing DHAR
36 were better protected from exposure to either chronic (100 ppb O3 4-h/day for 30 days) or
37 acute (200 ppb O3 for 2 hours) conditions than control plants and those with reduced
3 8 expression of DHAR (Chen and Gallie. 2005). The DHAR over-expressing plants
39 exhibited an increase in guard cell ascorbic acid, leading to a decrease in stomatal
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1 responsiveness to O3 and an increase in stomatal conductance and O3 uptake. Despite
2 this, the presence of higher levels of ascorbic acid led to a lower oxidative load and a
3 higher level of photosynthetic activity in the DHAR over-expressing plants (Chen and
4 Gallie. 2005). A subsequent study with tobacco plants over-expressing DHAR confirmed
5 some of these results. Levels of ascorbic acid were higher in the transgenic tobacco
6 plants, and they exhibited greater tolerance to O3 exposure (200 ppb O3) as demonstrated
7 by higher photosynthetic rates in the transgenic plants as compared to the control plants
8 (Eltayeb et al.. 2006). Over-expression of monodehydroascorbate reductase (MDAR) in
9 tobacco plants also showed enhanced stress tolerance in response to O3 exposure
10 (200 ppb O3), with higher rates of photosynthesis and higher levels of reduced ascorbic
11 acid as compared to controls (Eltaveb et al., 2007). Results of these studies demonstrate
12 the importance of ascorbic acid as a detoxification mechanism in some plant species, and
13 also emphasize that the recycling of oxidized ascorbate to maintain a reduced pool of
14 ascorbate is a factor in determining plant tolerance to oxidative stress.
15 The roles of other antioxidant metabolites and enzymes, including GSH, catalase (CAT),
16 and superoxide dismutase (SOD), were comprehensively reviewed in the 2006 O3
17 AQCD. Additional studies have supported the findings reported in that document.
18 Superoxide dismutase (SOD) and peroxidase (POD) activities were measured in both the
19 tolerant Bel B and sensitive Bel W3 tobacco cultivars exposed to ambient O3
20 concentrations for 2 weeks 3 times throughout a growing season (Borowiak et al.. 2009).
21 In this study, SOD and POD activity, including that of several different isoforms,
22 increased in both the sensitive and tolerant tobacco cultivars with exposure to O3,
23 however the isoenzyme composition for POD differed between the sensitive and tolerant
24 tobacco cultivars (Borowiak et al.. 2009) Tulip poplar (Liriodendron tulipifera) trees
25 exposed to increasing O3 concentrations (from 100 to 300 ppb O3 during a 2-week
26 period) showed increases in activities of SOD, ascorbate peroxidase (APX), glutathione
27 reductase (GR), MDAR, DHAR, CAT and POD in the 2-week period, although
28 individual enzyme activities increased at different times during the 2-week period (Ryang
29 et al.. 2009).
30 Longer, chronic O3 exposures in trees revealed increases in SOD and APX activity in
31 Quercus mongolica after 45 days of plant exposure to 80 ppb O3, which were followed
32 by declines in the activities and quantities of these enzymes after 75 days of exposure
33 (Yanetal.. 2010). Similarly, activities of SOD, APX, DHAR, MDAR, and GR increased
34 in Gingko biloba trees during the first 50 days of exposure to 80 ppb O3, followed by
35 decreases in activity below control values after 50 days of exposure (He etal.. 2006).
36 Soybean plants exposed to 70 or 100 ppb O3 for 4-h/day over the course of a growing
37 season showed elevated POD activity and a decrease in CAT activity at 40 and 60 days
38 after germination (Singh et al., 2010a).
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1 Antioxidant enzymes and metabolites have been shown to play an important role in
2 determining plant tolerance to O3 and mediating plant responses to O3. However, there is
3 also some evidence to suggest that the direct reaction of ascorbate with O3 could lead to
4 the formation of secondary toxicants, such as peroxy compounds, which may act upon
5 signal transduction pathways and modulate plant response to O3 (Sandermann, 2008).
6 Therefore, the role of ascorbate and other antioxidants and their interaction with other
7 plant responses to O3, such as the activation of signal transduction pathways, is likely far
8 more complex than is currently understood.
9.3.5 Effects on Primary and Secondary Metabolism
9.3.5.1 Light and Dark Reactions of Photosynthesis
9 Declines in the rate of photosynthesis and stomatal conductance in O3-treated plants have
10 been documented for many different plant species (Booker et al., 2009; U.S. EPA. 2006b)
11 (Wittig et al.. 2007). The 2006 O3 AQCD outlined what is known about the effects of O3
12 on carbon assimilation, and the more recent scientific literature confirms these findings.
13 While several measures of the light reactions of photosynthesis are sensitive to exposure
14 to O3 (see below), photosynthetic carbon assimilation is generally considered to be more
15 affected by pollutant exposure, resulting in an overall decline in photosynthesis (Guidi
16 and Degl'lnnocenti. 2008; Heath. 2008; Fiscus et al.. 2005). Loss of carbon assimilation
17 capacity has been shown to result primarily from declines in the quantity of Rubisco
18 (Singh et al., 2009; Calatayud et al.. 2007a). Experimental evidence suggests that both
19 decreases in Rubisco synthesis and enhanced degradation of the protein contribute to the
20 measured reduction in its quantity (U.S. EPA, 2006b). Reduced carbon assimilation has
21 been linked to reductions in biomass and yield (Wang et al.. 2009b; He et al.. 2007;
22 Novak et al.. 2007; Gregg et al., 2006; Keutgen et al., 2005). Recent studies evaluating
23 O3 induced changes in the transcriptome and proteome of several different species
24 confirm these findings. Levels of mRNA for the small subunit of Rubisco (rbcS) declined
25 in European beech saplings exposed to 300 ppb O3 for 8-h/day for up to 26 days (Olbrich
26 et al., 2005). Similar declines in rbcS mRNA were also measured in the beech saplings in
27 a free air exposure system over a course of two growing seasons (Olbrich et al.. 2009).
28 Proteomics studies have also confirmed the effects of O3 on proteins involved in carbon
29 assimilation. Reductions in quantities of the small and large subunit (rbcL) of Rubisco
30 and Rubisco activase were measured in soybean plants exposed to 120 ppb O3 for 3 days
31 in growth chambers (Ahsan et al.. 2010). Exposure of young poplar trees to 120 ppb O3
32 for 35 days in exposure chambers resulted in reductions of Rubisco, Rubisco activase,
33 and up to 24 isoforms of Calvin cycle enzymes, most of which play a role in regenerating
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1 the CO2 acceptor molecule, ribulose-1.5-bisphosphate (Bohler et al., 2007). Reductions
2 in protein quantity of both the small and large subunit of Rubisco were seen in wheat
3 plants exposed to ambient (average concentration 47.3 ppb O3) and elevated O3 (ambient
4 + 10 or 20 ppb O3) in open-top chambers for 5-h/day for 50 days (Sarkar et al.. 2010).
5 Lettuce plants exposed to 100 ppb O3 in growth chambers for 8-h/day for 3 weeks also
6 showed reductions in transcript and protein levels of the small and large subunits of
7 Rubisco and Rubisco activase (Goumenaki et al.. 2010). The reductions in carbon
8 assimilation have been associated with declines in both the mRNA of the small and large
9 subunits of Rubisco, and with reductions in Rubisco activase mRNA and protein.
10 Additionally, the reduction in Rubisco quantity has also been associated with the O3-
11 induced oxidative modification of the enzyme, which is evidenced by the increases in
12 carbonyl groups on the protein after plant exposure to O3.
13 In addition to impacts on carbon assimilation, the deleterious effects of O3 on the
14 photosynthetic light reactions have received more attention in recent years. Chlorophyll
15 fluorescence provides a useful measure of changes to the photosynthetic process from
16 exposure to oxidative stress. Decreases in the Fv/Fm ratio (a measure of the maximum
17 efficiency of Photosystem II) in dark adapted leaves indicate a decline in the efficiency of
18 the PSII photosystems and a concomitant increase in non-photochemical quenching
19 (Guidi and Degl'lnnocenti. 2008; Scebbaetal.. 2006). Changes in these parameters have
20 been correlated to differential sensitivity of plants to the pollutant. In a study to evaluate
21 the response of 4 maple species to O3 (exposed to an 8-h avg of 51 ppb for ambient and
22 79 ppb for elevated treatment in OTC), the 2 species which were most sensitive based on
23 visible injury and declines in CO2 assimilation also showed the greatest decreases in
24 Fv/Fm in symptomatic leaves. In asymptomatic leaves, CO2 assimilation decreased
25 significantly but there was no significant decline in Fv/Fm (Calatavud et al., 2007a). Degl
26 'Innocenti et al. (2007) measured significant decreases in Fv/Fm in young and
27 symptomatic leaves of a resistant tomato genotype (line 93.1033/1) in response to O3
28 exposure (150 ppb O3 for 3 hours in a growth chamber), but only minor decreases in
29 asymptomatic leaves with no associated changes in net photosynthetic rate. In the O3
30 sensitive tomato cultivar Cuor Di Bue, the Fv/Fm ratio did not change, while the
31 photosynthetic rate declined significantly in asymptomatic leaves (Degl'Innocenti et al..
32 2007). In two soybean cultivars, Fv/Fm also declined significantly with plant exposure to
33 O3 (Singh et al., 2009). It appears that in asymptomatic leaves, photoinhibition, as
34 indicated by a decrease in Fv/Fm, is not the main reason for a decline in photosynthesis.
35 An evaluation of photosynthetic parameters of two white clover (Trifolium repens cv.
36 Regal) clones that differ in their O3 sensitivity revealed that O3 (40-110 ppb O3 for 7-
37 h/day for 5 days) increased the coefficient of non-photochemical quenching (QM>) in both
38 the resistant (NC-R) and sensitive (NC-S) clones, however q^p was significantly lower
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1 for the sensitive clone (Crous et al., 2006). Sensitive Acer clones had a lower coefficient
2 of non-photochemical quenching, while exposure to O3 increased q^p in both sensitive
3 and tolerant clones (Calatavud et al., 2007a). While exposure to O3 also increased q^p in
4 tomato, there were no differences in the coefficient of photochemical quenching between
5 cultivars thought to be differentially sensitive to O3. (Degl'Innocenti et al., 2007). Higher
6 qM. as a result of exposure to O3 indicates a reduction in the proportion of absorbed light
7 energy being used to drive photochemistry. A lower coefficient of non-photochemical
8 quenching in O3 sensitive plants could indicate increased vulnerability to ROS generated
9 during exposure to oxidative stress (Crous et al.. 2006).
10 Most of the research on O3 effects on photosynthesis has focused on C3 (Calvin cycle)
11 plants because C4 (Hatch-Slack) plants have lower stomatal conductance and are,
12 therefore, thought to be less sensitive to O3 stress. However, a few studies have been
13 conducted to evaluate the effects of O3 on C4 photosynthesis. In older maize leaves,
14 Leitao et al. (2007b; 2007c) found that the activity, quantity and transcript levels of both
15 Rubisco and phosphoenolpyruvate carboxylase (PEPc) decreased as a function of rising
16 O3 concentration. In younger maize leaves, the quantity, activity, and transcript levels of
17 the carboxylases were either increased or unaffected in plants exposed to 40 ppb O3 for
18 7- h/day for 28-33 days, but decreased at 80 ppb (Leitao et al.. 2007a: Leitao et al..
19 2007b).
9.3.5.2 Respiration and Dark Respiration
20 While much research emphasis regarding O3 effects on plants has focused on the
21 negative impacts on carbon assimilation, other studies have measured impacts on
22 catabolic pathways such as shoot respiration and photorespiration. Generally, shoot
23 respiration has been found to increase in plants exposed to O3. Bean plants exposed to
24 ambient (average 12-h mean 43 ppb) and twice ambient (average 12-h mean 80 ppb) O3
25 showed increases in respiration. When mathematically partitioned, the maintenance
26 coefficient of respiration was significantly increased in O3 treated plants, while the
27 growth coefficient of respiration was not affected (Amthor. 1988). Loblolly pines were
28 exposed to ambient (12-h daily mean was 45 ppb) and twice ambient (12 hours daily
29 mean was 86 ppb) O3 for 12-h/day for approximately seven months per year for 3 and
30 4 years. While photosynthetic activity declined with the age of the needles and increasing
31 O3 concentration, enzymes associated with respiration showed higher levels of activity
32 with increasing O3 concentration (Dizengremel et al., 1994). In their review on the role of
33 metabolic changes in plant redox status after O3 exposure, Dizengremel et al. (2009)
34 summarized multiple studies in which several different tree species were exposed to O3
35 concentrations ranging from ambient to 200 ppb O3 for at least several weeks. In all
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1 cases, the activity of enzymes, including phosphofructokinase, pyruvate kinase and
2 fumarase, which are part of several catabolic pathways, were increased in O3 treated
3 plants.
4 Photorespiration is a light-stimulated process which consumes O2 and releases CO2.
5 While it has been regarded as a wasteful process, more recent evidence suggests that it
6 may play a role in photoprotection during photosynthesis (Bagard et al., 2008). The few
7 studies that have been conducted on O3 effects on photorespiration suggest that rates of
8 photorespiration decline concomitantly with rates of photosynthesis. Soybean plants were
9 exposed to ambient (daily averages 43-58 ppb) and 1.5 ambient O3 (daily averages 63-
10 83 ppb) O3 in OTCs for 12-h/day for 4 months. Rates of photosynthesis and
11 photorespiration and photorespiratory enzyme activity declined only at the end of the
12 growing season and did not appear to be very sensitive to O3 exposure (Booker et al.,
13 1997). Young hybrid poplars exposed to 120 ppb O3 for 13-h/day for 35 days in
14 phytotron chambers showed that effects on photorespiration and photosynthesis were
15 dependent upon the developmental stage of the leaf. While young leaves were not
16 impacted, reductions in photosynthesis and photorespiration were measured in fully
17 expanded leaves (Bagard et al.. 2008).
9.3.5.3 Secondary Metabolism
18 Transcriptome analysis of Arabidopsis plants has revealed modulation of several genes
19 involved in plant secondary metabolism (Ludwikow and Sadowski, 2008). Phenylalanine
20 ammonia lyase (PAL) has been the focus of many studies involving plant exposure to O3
21 due to its importance in linking the phenylpropanoid pathway of plant secondary
22 metabolism to primary metabolism in the form of the shikimate pathway. Genes encoding
23 several enzymes of the phenylpropanoid pathway and lignin biosynthesis were up-
24 regulated in transcriptome analysis of Arabidopsis plants (Col-0) exposed to 350 ppb O3
25 for 6 hours, while 2 genes involved in flavonoid biosynthesis were down-regulated
26 (Ludwikow et al., 2004). Exposure of Arabidopsis (Col-0) to lower O3 concentrations
27 (150 ppb for 8-h/day for 2 days) resulted in the induction of 11 transcripts involved in
28 flavonoid synthesis. In their exposure of 2-year-old Mediterranean shrub Phillyrea
29 latifolia to 110 ppb O3 for 90 days, Paolacci et al. (2007) identified four clones that were
30 up-regulated and corresponded to genes involved in the synthesis of secondary
31 metabolites, such as isoprenoids, polyamines and phenylpropanoids. Up-regulation of
32 genes involved in isoprene synthesis was also observed mMedicago trunculata exposed
33 to 300 ppb O3 for 6 hours, while genes encoding enzymes of the flavonoid synthesis
34 pathway were either up- or down-regulated (Puckette et al., 2008). Exposure of red clover
35 to 1.5 x ambient O3 (average concentrations of 32.4 ppb) for up to 9 weeks in an open
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1 field exposure system resulted in increases in leaf total phenolic content. However, the
2 types of phenolics that were increased in response to O3 exposure differed depending
3 upon the developmental stage of the plant. While almost all of the 31 different phenolic
4 compounds measured increased in quantity initially during the exposure, after 3 weeks
5 the quantity of isoflavones decreased while other phenolics increased (Saviranta et al.,
6 2010). Exposure of beech saplings to ambient and 2 x ambient O3 concentrations over 2
7 growing seasons resulted in the induction of several enzymes which contribute to lignin
8 formation, while enzymes involved in flavonoid biosynthesis were down-regulated
9 (Olbrich et al., 2009). Exposure of tobacco Bel W3 to 160 ppb O3 for 5 hours showed up-
10 regulation of almost all genes encoding for enzymes which are part of the prechorismate
11 pathway (Janzik et al., 2005). Isoprenoids can serve as antioxidant compounds in plants
12 exposed to oxidative stress (Paolacci et al.. 2007).
13 The prechorismate pathway is the pathway leading to the formation of chorismate, a
14 precursor to the formation of the aromatic amino acids tryptophan, tyrosine and
15 phenylalanine. These amino acids are precursors for the formation of many secondary
16 aromatic compounds, and, therefore, the prechorismate pathway represents a branch-
17 point in the regulation of metabolites into either primary or secondary metabolism (Janzik
18 et al.. 2005). Exposure of the O3 sensitive Bel W3 tobacco cultivar at 160 ppb for 5 hours
19 showed an increase in transcript levels of most of the genes encoding enzymes of the
20 prechorismate pathway. However, shikimate kinase (SK) did not show any change in
21 transcript levels and only one of three isoforms of DAHPS (3-deoxy-D-arabino-
22 heptulosonat-7-phosphate synthase), the first enzyme in this pathway, was induced by O3
23 exposure (Janzik et al., 2005). Differential induction of DAHPS isoforms was also
24 observed in European beech after 40 days of exposure to 150-190 ppb O3. At this time
25 point in the beech experiment, transcript levels of shikimate pathway enzymes, including
26 SK, were generally strongly induced after an only weak initial induction after the first
27 40 days of exposure. Both soluble and cell-wall bound phenolic metabolites showed only
28 minimal increases in response to O3 for the duration of the exposure period (Alonso et
29 al., 2007). Total leaf phenolics decreased with leafage in Populus nigra exposed to
30 80 ppb O3 for 12-h/day for 14 days. Ozone increased the concentration of total leaf
31 phenolics in newly expanded leaves, with the most significant increases occurring in
32 compounds such as quercitin glycoside, which has a high antioxidant capacity (Fares et
33 al., 201 Ob). While several phenylpropanoid pathway enzymes were induced in two poplar
34 clones exposed to 60 ppb O3 for 5-h/day for 15 days, the degree of induction differed
35 between the two clones. In the tolerant 1-214 clone, PAL activity increased nine fold in
36 O3-treated plants as compared to controls, while there was no significant difference in
37 PAL activity in the sensitive Eridano clone (Di Baccio et al.. 2008).
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1 Polyamines such as putrescine, spermidine and spermine play a variety of roles in plants
2 and have been implicated in plant defense responses to both abiotic and biotic stresses.
3 They exist in both a free form and conjugated to hydroxycinnamic acids. Investigations
4 on the role of polyamines have found that levels of putrescine increase in response to
5 oxidative stress. This increase stems largely from the increase in the activity of arginine
6 decarboxylase (ADC), a key enzyme in the synthesis of putrescine (Groppa and
7 Benavides. 2008). Langebartels et al. (1991) described differences in putrescine
8 accumulation in O3-treated tobacco plants exposed to several O3 concentrations, ranging
9 from 0-400 ppb for 5-7 hours. A large and rapid increase in putrescine occurred in the
10 tolerant Bel B cultivar and only a small increase in the sensitive Bel W3 cultivar, which
11 occurred only after the formation of necrotic leaf lesions. Van Buuren et al. (2002)
12 further examined the role of polyamines in these two tobacco cultivars during an acute
13 (130 ppb O3 for 7-h in a growth chamber) exposure. They found that while free
14 putrescine accumulated in undamaged tissue of both cultivars, conjugated putrescine
15 predominantly accumulated in tissues undergoing cell death after plant exposure to O3
16 (van Buuren et al.. 2002). The authors suggest that while free putrescine may not play a
17 role in conferring tolerance in the Bel B cultivar, conjugated putrescine may play a role in
18 O3-induced programmed cell death in Bel W3 plants.
19 Isoprene is emitted by some plant species and represents the predominant biogenic source
20 of hydrocarbon emissions in the atmosphere (Guenther et al.. 2006). In the atmosphere,
21 the oxidation of isoprene by hydroxyl radicals can enhance O3 formation in the presence
22 of NOX, thereby impacting the O3 concentration that plants are exposed to. While
23 isoprene emission varies widely between species, it has been proposed to stabilize
24 membranes and provide those plant species that produce it with a mechanism of
25 thermotolerance (Sharkey et al.. 2008). It has also been suggested that isoprene may act
26 as an antioxidant compound to scavenge O3 (Loreto and Velikova. 2001). Recent studies
27 using a variety of plant species have shown conflicting results in trying to understand the
28 effects of O3 on isoprene emission. Exposure to acute doses of O3 (300 ppb for 3-h) in
29 detached leaves ofPhmgmites australis resulted in stimulation of isoprene emissions
30 (Velikova et al.. 2005). Similar increases in isoprene emissions were measured in
31 Populus nigra after exposure to 100 ppb O3 for 5 days continuously (Fares etal., 2008).
32 Isoprene emission in attached leaves of Populus alba, which were exposed to 150 ppb O3
33 for 11-h/day for 30 days inside cuvettes, was inhibited, while isoprene emission and
34 transcript levels of isoprene synthase mRNA were increased in the leaves exposed to
35 ambient O3 (40 ppb), which were located above the leaves enclosed in the exposure
36 cuvettes (Fares et al.. 2006). Exposure of 2 genotypes of hybrid poplar to 120 ppb O3 for
37 6-h/day for 8 days resulted in a significant reduction in isoprene emission in the O3-
38 sensitive but not the tolerant genotype (Ryan et al.. 2009). Similarly, O3 treatment
39 (80 ppb 12-h/day for 14 days) of Populus nigra showed that isoprene emission was
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1 reduced in the treated plants relative to the control plants (Fares et al., 201 Ob). Based on
2 results of this and other studies, Fares et al. (201 Ob) concluded that the isoprenoid
3 pathway may be induced in plants exposed to acute O3 doses, while at lower doses
4 isoprene emission may be inhibited. Vickers et al. (2009) developed transgenic tobacco
5 plants with the isoprene synthase gene from Populus alba and exposed them to 120 ppb
6 O3 for 6-h/day for 2 days. They determined that the wildtype plants showed significantly
7 more O3 damage, including the development of leaf lesions and a decline in
8 photosynthetic rates, than the transgenic, isoprene-emitting plants. Transgenic plants also
9 accumulated less H2O2 and had lower levels of lipid peroxidation following exposure to
10 O3 than the wildtype plants (Vickers et al.. 2009). These results indicate that isoprene
11 may have a protective role for plants exposed to oxidative stress.
9.3.6 Summary
12 The results of recent studies on the effects of O3 stress on plants support and strengthen
13 those reported in the 2006 O3 AQCD. The most significant new body of evidence since
14 the 2006 O3 AQCD comes from research on molecular mechanisms of the biochemical
15 and physiological changes observed in many plant species in response to O3 exposure.
16 Recent studies have employed new techniques, such as those used in evaluating
17 transcriptomes and proteomes to perform very comprehensive analyses of changes in
18 gene transcription and protein expression in plants exposed to O3. These newer molecular
19 studies not only provide very important information regarding the many mechanisms of
20 plant responses to O3, they also allow for the analysis of interactions between various
21 biochemical pathways which are induced in response to O3. However, many of these
22 studies have been conducted in artificial conditions with model plants, which are
23 typically exposed to very high, short doses of O3. Therefore, additional work remains to
24 elucidate whether these plant responses are transferable to other plant species exposed to
25 more realistic ambient conditions.
26 Ozone is taken up into leaves through open stomata. Once inside the substomatal cavity,
27 O3 is thought to rapidly react with the aqueous layer surrounding the cell (apoplast) to
28 form breakdown products such as hydrogen peroxide (H2O2), superoxide (O2"), hydroxyl
29 radicals (HO) and peroxy radicals (HO2). Plants could be sensing the presence of O3
30 and/or its breakdown products in a variety different ways, depending upon the plant
31 species and the exposure parameters. Experimental evidence suggests that mitogen-
32 activated protein kinases and calcium are important components of the signal
33 transduction pathways, which communicate signals to the nucleus and lead to changes in
34 gene expression in response to O3. It is probable that there are multiple sensing
3 5 mechanisms and signal transduction pathways, and their activation may depend upon the
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1 plant species, its developmental stage and/or O3 exposure conditions. Initiation of signal
2 transduction pathways in O3 treated plants has also been observed in stomatal guard cells.
3 Reductions in stomatal conductance in have been described for many plant species
4 exposed to O3, and new experimental evidence suggests that this reduction may be due
5 not only to a decrease in carboxylation efficiency, but also to a direct impact of O3 on
6 stomatal guard cell function, leading to a changes in stomatal conductance.
7 Alterations in gene transcription that have been observed in O3-treated plants are now
8 evaluated more comprehensively using DNA microarray studies, which measure changes
9 in the entire transcriptome rather than measuring the transcript levels of individual genes.
10 These studies have demonstrated very consistent trends, even though O3 exposure
11 conditions (concentration, duration of exposure), plant species and sampling times vary
12 significantly. Genes involved in plant defense, signaling, and those associated with the
13 synthesis of plant hormones and secondary metabolism are generally up-regulated in
14 plants exposed to O3, while those related to photosynthesis and general metabolism are
15 typically down-regulated. Proteome studies support these results by demonstrating
16 concomitant increases or decreases in the proteins encoded by these genes. Transcriptome
17 analysis has also illuminated the complex interactions that exist between several different
18 phytohormones and how they modulate plant sensitivity and response to O3.
19 Experimental evidence suggests that while ethylene and salicylic acid are needed to
20 develop O3-induced leaf lesions, jasmonic acid acts antagonistically to ethylene and
21 salicylic acid to limit the spread of the lesions. Abscisic acid, in addition to its role in
22 regulating stomatal aperture, may also act antagonistically to the jasmonic acid signaling
23 pathway. Changes in the quantity and activity of these phytohormones and the
24 interactions between them reveal some of the complexity of plant responses to an
25 oxidative stressor such as O3.
26 Another critical area of interest is to better understand and quantify the capacity of the
27 plant to detoxify oxygen radicals using antioxidant metabolites, such as ascorbate and
28 glutathione, and the enzymes that regenerate them. Ascorbate remains an important focus
29 of research, and, due to its location in the apoplast in addition to other cellular
30 compartments, it is regarded as a first line of defense against oxygen radicals formed in
31 the apoplast. Most studies demonstrate that antioxidant metabolites and enzymes increase
32 in quantity and activity in plants exposed to O3, indicating that they play an important
33 role in protecting plants from oxidative stress. However, attempts to quantify the
34 detoxification capacity of plants have remained unsuccessful, as high quantities of
3 5 antioxidant metabolites and enzymes do not always translate into greater protection of the
36 plant. Considerable variation exists between plant species, different developmental
37 stages, and the environmental and O3 exposure conditions which plants are exposed to.
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1 As indicated earlier, the described alterations in transcript levels of genes correlate with
2 observed changes quantity and activity of the enzymes and metabolites involved in
3 primary and secondary metabolism. In addition to the generalized up-regulation of the
4 antioxidant defense system, photosynthesis typically declines in O3 treated plants.
5 Declines in C fixation due to reductions in quantity and activity of Rubisco were
6 extensively described in the 2006 O3 AQCD. More recent studies support these results
7 and indicate that declines in Rubisco activity may also result from reductions in Rubisco
8 activase enzyme quantity. Other studies, which have focused on the light reactions of
9 photosynthesis, demonstrate that plant exposure to O3 results in declines in electron
10 transport efficiency and a decreased capacity to quench oxidizing radicals. Therefore, the
11 overall declines in photosynthesis observed in O3 -treated plants likely result from
12 combined impacts on stomatal conductance, carbon fixation and the light reactions.
13 While photosynthesis generally declines in plants exposed to O3, catabolic pathways such
14 as respiration have been shown to increase. It has been hypothesized that increased
15 respiration may result from greater energy needs for defense and repair. Secondary
16 metabolism is generally up-regulated in a variety of species exposed to O3 as a part of a
17 generalized plant defense mechanism. Some secondary metabolites, such as flavonoids
18 and polyamines, are of particular interest as they are known to have antioxidant
19 properties. The combination of decreases in C assimilation and increases in catabolism
20 and the production of secondary metabolites would negatively impact plants by
21 decreasing the energy available for growth and reproduction.
9.4 Nature of Effects on Vegetation and Ecosystems
9.4.1 Introduction
22 Ambient O3 concentrations have long been known to cause visible symptoms, decreases
23 in photosynthetic rates, decreases in growth and yield of plants as well as many other
24 effects on ecosystems (U.S. EPA. 2006b. 1996c. 1986. 1978a). Numerous studies have
25 related O3 exposure to plant responses, with most effort focused on the yield of crops and
26 the growth of tree seedlings. Many experiments exposed individual plants grown in pots
27 or soil under controlled conditions to known concentrations of O3 for a segment of
28 daylight hours for some portion of the plant's life span. Information in this section also
29 goes beyond individual plant scale responses to consider effects at the broader ecosystem
30 scale, including effects related to ecosystem services.
31 This section will focus mainly on studies published since the release of the 2006 O3
32 AQCD. However, because much O3 research was conducted prior to the 2006 O3 AQCD,
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1 the present discussion of vegetation and ecosystem response to O3 exposure is largely
2 based on the conclusions of the 1978, 1986, 1996, and 2006 O3 AQCDs.
9.4.1.1 Ecosystem Scale, Function, and Structure
3 Information presented in this section was collected at multiple spatial scales, ranging
4 from the physiology of a given species to population, community, and ecosystem
5 investigations. An ecological population is a group of individuals of the same species and
6 a community is an assemblage of populations of different species interacting with one
7 another that inhabit an area. For this assessment, "ecosystem" is defined as the interactive
8 system formed from all living organisms and their abiotic (physical and chemical)
9 environment within a given area (IPCC. 2007a). The boundaries of what could be called
10 an ecosystem are somewhat arbitrary, depending on the focus of interest or study. Thus,
11 the extent of an ecosystem may range from very small spatial scales to, ultimately, the
12 entire Earth (IPCC. 2007a). All ecosystems, regardless of size or complexity, have
13 interactions and physical exchanges between biota and abiotic factors, this includes both
14 structural (e.g., soil type and food web trophic levels) and functional (e.g., energy flow,
15 decomposition, nitrification) attributes.
16 Ecosystems are most often defined by their structure based on the number and type of
17 species present. Structure may refer to a variety of measurements including the species
18 richness, abundance, community composition and biodiversity as well as landscape
19 attributes. Competition among and within species and their tolerance to environmental
20 stressors are key elements of survivorship. When environmental conditions are shifted,
21 for example, by the presence of anthropogenic air pollution, these competitive
22 relationships may change and tolerance to stress may be exceeded. Ecosystems may also
23 be defined on a functional basis. "Function" refers to the suite of processes and
24 interactions among the ecosystem components and their environment that involve
25 nutrient and energy flow as well as other attributes including water dynamics and the flux
26 of trace gases. Plant processes including photosynthesis, respiration, C allocation,
27 nutrient uptake and evaporation, are directly related to functions of energy flow and C,
28 nutrient and water cycling. The energy accumulated and stored by vegetation (via
29 photosynthetic C capture) is available to other organisms. Energy moves from one
30 organism to another through food webs, until it is ultimately released as heat. Nutrients
31 and water can be recycled. Air pollution alters the function of ecosystems when elemental
32 cycles or the energy flow are altered. This alteration can also be manifested in changes in
33 the biotic composition of ecosystems.
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1 There are at least three levels of ecosystem response to pollutants: (1) the individual
2 organism and its environment; (2) the population and its environment; and (3) the
3 biological community composed of many species and their environment (Billings. 1978).
4 Individual organisms within a population vary in their ability to withstand the stress of
5 environmental change. The response of individual organisms within a population is based
6 on their genetic constitution, stage of growth at time of exposure to stress, and the
7 microhabitat in which they are growing (Levine and Pinto. 1998). The stress range within
8 which organisms can exist and function determines the ability of the population to
9 survive. Those best able to cope with environmental stressors survive and reproduce.
10 Competition among different species results in succession (community change over time)
11 and, ultimately, sensitive species may be progressively replaced and communities shift to
12 favor those species that may have the capability to tolerate stressors such as O3 (Rapport
13 and Whitford. 1999; Guderian. 1985).
9.4.1.2 Ecosystem Services
14 Ecosystem structure and function may be translated into ecosystem services. Ecosystem
15 services are the benefits people obtain from ecosystems (UNEP. 2003). Ecosystems
16 provide many goods and services that are of vital importance for the functioning of the
17 biosphere and provide the basis for the delivery of tangible benefits to human society.
18 Hassan et al. (2005) define these benefits to include supporting, provisioning, regulating,
19 and cultural services:
20 • Supporting services are necessary for the production of all other ecosystem
21 services. Some examples include biomass production, production of
22 atmospheric O2, soil formation and retention, nutrient cycling, water cycling,
23 and provisioning of habitat. Biodiversity is a supporting service that is
24 increasingly recognized to sustain many of the goods and services that humans
25 enjoy from ecosystems. These provide a basis for three higher-level categories
26 of services.
27 • Provisioning services, such as products (Gitay et al.. 2001). i.e., food
28 (including game, roots, seeds, nuts and other fruit, spices, fodder), water, fiber
29 (including wood, textiles), and medicinal and cosmetic products (such as
30 aromatic plants, pigments).
31 • Regulating services that are of paramount importance for human society such
32 as (1) C sequestration, (2) climate and water regulation, (3) protection from
33 natural hazards such as floods, avalanches, or rock-fall, (4) water and air
34 purification, and (5) disease and pest regulation.
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1 • Cultural services that satisfy human spiritual and aesthetic appreciation of
2 ecosystems and their components including recreational and other nonmaterial
3 benefits.
4 In the sections that follow, available information on individual, population and
5 community response to O3 will be discussed. Effects of O3 on productivity and
6 C sequestration, water cycling, below-ground processes, competition and biodiversity,
7 and insects and wildlife are considered below and in the context of ecosystem services
8 where appropriate.
9.4.2 Visible Foliar Injury and Biomonitoring
9 Visible foliar injury resulting from exposure to O3 has been well characterized and
10 documented over several decades on many tree, shrub, herbaceous, and crop species
11 (U.S. EPA. 2006b. 1996b. 1984. 1978a). Visible foliar injury symptoms are considered
12 diagnostic as they have been verified experimentally in exposure-response studies, using
13 exposure methodologies such as CSTRs, OTCs, and free-air fumigation (see Section 9.2
14 for more detail on exposure methodologies). Several pictorial atlases and guides have
15 been published, providing details on diagnosis and identification of O3-induced visible
16 foliar injury on many plant species throughout North America (Flagler. 1998; NAPAP.
17 1987) and Europe (Innes etal.. 2001; Sanchez et al.. 2001). Typical visible injury
18 symptoms on broad-leaved plants include: stippling, flecking, surface bleaching, bifacial
19 necrosis, pigmentation (e.g., bronzing), chlorosis, and/or premature senescence. Typical
20 visible injury symptoms for conifers include: chlorotic banding, tip burn, flecking,
21 chlorotic mottling, and/or premature senescence of needles. Although common patterns
22 of injury develop within a species, these foliar lesions can vary considerably between and
23 within taxonomic groups. Furthermore, the degree and extent of visible foliar injury
24 development varies from year to year and site to site (Orendovici-Best et al.. 2008;
25 Chappelka et al.. 2007; Smith et al.. 2003). even among co-members of a population
26 exposed to similar O3 levels, due to the influence of co-occurring environmental and
27 genetic factors. Nevertheless, Chappelka et al. (2007) reported that the average incidence
28 of O3-induced foliar injury was 73% on milkweed in the Great Smoky Mountains
29 National Park in the years 1992-1996.
30 Although the majority of O3-induced visible foliar injury occurrence has been observed
31 on seedlings and small plants, many studies have reported visible injury of mature
32 coniferous trees, primarily in the western U.S. (Arbaugh et al.. 1998) and to mature
33 deciduous trees in eastern North America (Schaub et al.. 2005; Vollenweider et al.. 2003;
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1 Chappelka et al.. 1999a: Chappelka et al.. 1999b: Somerset al.. 1998; Hildebrand et al.
2 1996V
3 It is important to note that visible foliar injury occurs only when sensitive plants are
4 exposed to elevated O3 concentrations in a predisposing environment. A major modifying
5 factor for O3-induced visible foliar injury is the amount of soil moisture available to a
6 plant during the year that the visible foliar injury is being assessed. This is because lack
7 of soil moisture generally decreases stomatal conductance of plants and, therefore, limits
8 the amount of O3 entering the leaf that can cause injury (Matvssek et al.. 2006; Panek.
9 2004: Grulke et al.. 2003a: Panek and Goldstein. 2001: Temple et al.. 1992: Temple et
10 al., 1988). Consequently, many studies have shown that dry periods in local areas tend to
11 decrease the incidence and severity of O3-induced visible foliar injury; therefore, the
12 incidence of visible foliar injury is not always higher in years and areas with higher O3,
13 especially with co-occurring drought (Smith et al.. 2003). Other factors such as leafage
14 influence the severity of symptom expression with older leaves showing greater injury
15 severity as a result of greater seasonal exposure (Zhang et al.. 2010a).
16 Although visible injury is a valuable indicator of the presence of phytotoxic
17 concentrations of O3 in ambient air, it is not always a reliable indicator of other negative
18 effects on vegetation. The significance of O3 injury at the leaf and whole plant levels
19 depends on how much of the total leaf area of the plant has been affected, as well as the
20 plant's age, size, developmental stage, and degree of functional redundancy among the
21 existing leaf area. Previous O3 AQCDs have noted the difficulty in relating visible foliar
22 injury symptoms to other vegetation effects such as individual plant growth, stand
23 growth, or ecosystem characteristics (U.S. EPA. 2006b. 1996b). As a result, it is not
24 presently possible to determine, with consistency across species and environments, what
25 degree of injury at the leaf level has significance to the vigor of the whole plant.
26 However, in some cases, visible foliar symptoms have been correlated with decreased
27 vegetative growth (Somers et al.. 1998: Karnosky et al.. 1996: Peterson etal.. 1987:
28 BenoitetaL 1982) and with impaired reproductive function (Chappelka. 2002: Black et
29 al.. 2000). Conversely, the lack of visible injury does not always indicate a lack of
30 phytotoxic concentrations of O3 or a lack of non-visible O3 effects (Gregg et al., 2006.
31 2003).
9.4.2.1 Biomonitoring
32 The use of biological indicators to detect phytotoxic levels of O3 is a longstanding and
33 effective methodology (Chappelka and Samuelson. 1998: Manning and Krupa. 1992). A
34 plant bioindicator can be defined as a vascular or nonvascular plant exhibiting a typical
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1 and verifiable response when exposed to a plant stress such as an air pollutant (Manning.
2 2003). To be considered a good indicator species, plants must (1) exhibit a distinct,
3 verified response; (2) have few or no confounding disease or pest problems; and (3)
4 exhibit genetic stability (U.S. EPA. 2006b). Such sensitive plants can be used to detect
5 the presence of a specific air pollutant such as O3 in the ambient air at a specific location
6 or region and, as a result of the magnitude of their response, provide unique information
7 regarding specific ambient air quality. Bioindicators can be either introduced sentinels,
8 such as the widely used tobacco (Nicotiana tabacuni) variety Bel W3 (Calatayud et al..
9 2007b: Laffrav et al.. 2007; Nali et al.. 2007; Gombert et al.. 2006; Kostka-Rick and
10 Hahn. 2005; Heggestad. 1991) or detectors, which are sensitive native plant species
11 (Chappelka et al.. 2007; Souza et al.. 2006). The approach is especially useful in areas
12 where O3 monitors are not operated (Manning. 2003). For example, in remote wilderness
13 areas where instrument monitoring is generally not available, the use of bioindicator
14 surveys in conjunction with the use of passive samplers (Krupa et al.. 2001) may be a
15 useful methodology (Manning. 2003). However, it requires expertise in recognizing those
16 signs and symptoms uniquely attributable to exposure to O3 as well as in their
17 quantitative assessment.
18 Since the 2006 O3 AQCD, new sensitive plant species have been identified from field
19 surveys and verified in controlled exposure studies (Kline et al.. 2009; Kline et al.. 2008).
20 Several multiple-year field surveys have also been conducted at National Wildlife
21 Refuges in Maine, Michigan, New Jersey, and South Carolina (Davis. 2009. 2007a. b;
22 Davis and Orendovici. 2006).
23 The USDA Forest Service through the Forest Health Monitoring Program (FHM) (1990 -
24 2001) and currently the Forest Inventory and Analysis (FIA) Program has been collecting
25 data regarding the incidence and severity of visible foliar injury on a variety of O3
26 sensitive plant species throughout the U.S. (Coulston et al.. 2003; Smith et al.. 2003). The
27 plots where these data are taken are known as biosites. These biosites are located
28 throughout the country and analysis of visible foliar injury within these sites follows a set
29 of established protocols. For more details, see http://www.nrs.fs.fed.us/fia/topics/ozone/
30 (USDA. 2011). The network has provided evidence of O3 concentrations high enough to
31 induce visible symptoms on sensitive vegetation. From repeated observations and
32 measurements made over a number of years, specific patterns of areas experiencing
33 visible O3 injury symptoms can be identified. Coulston et al. (2003) used information
34 gathered over a 6-year period (1994-1999) from the network to identify several species
35 that were sensitive to O3 over a regional scale including sweetgum (Liquidambar
36 styraciflud), loblolly pine (Pinus taedd), and black cherry (Prunus serotind). In a study of
37 the west coast of the U.S, Campbell et al. (2007) reported O3 injury in 25-37% of biosites
38 in California forested ecosystems from 2000-2005.
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1 A study by Kohut (2007) assessed the risk of O3-induced visible foliar injury on
2 bioindicator plants (NPS. 2006) in 244 national parks in support of the National Park
3 Service's Vital Signs Monitoring Network (NFS. 2007). The risk assessment was based
4 on a simple model relating response to the interaction of species, level of O3 exposure,
5 and exposure environment. Kohut (2007) concluded that the risk of visible foliar injury
6 was high in 65 parks (27%), moderate in 46 parks (19%), and low in 131 parks (54%).
7 Some of the well-known parks with a high risk of O3-induced visible foliar injury include
8 Gettysburg, Valley Forge, Delaware Water Gap, Cape Cod, Fire Island, Antietam,
9 Harpers Ferry, Manassas, Wolf Trap Farm Park, Mammoth Cave, Shiloh, Sleeping Bear
10 Dunes, Great Smoky Mountains, Joshua Tree, Sequoia and Kings Canyon, and Yosemite.
11 Lichens have also long been used as biomonitors of air pollution effects on forest health
12 (Nash. 2008). It has been suspected, based on field surveys in the San Bernardino
13 Mountains surrounding the Los Angeles air basin, that declines in lichen diversity and
14 abundance were correlated with measured O3 gradients (Gul et al., 2011). Several recent
15 studies in North America (Geiser and Neitlich. 2007; Gombert et al.. 2006; Jovan and
16 McCune. 2006) and Europe (Nali et al.. 2007; Gombert et al., 2006) have used lichens as
17 biomonitors of atmospheric deposition (e.g., N and S) and O3 exposure. Nali et al. (2007)
18 found that epiphytic lichen biodiversity was not related to O3 geographical distribution.
19 In addition, a recent study by Riddell et al. (2010) found that lichen species, Ramalina
20 menziesii, showed no decline in physiological response to low and moderate
21 concentrations of O3 and may not be a good indicator for O3 pollution. Mosses have also
22 been used as biomonitors of air pollution; however, there remains a knowledge gap in the
23 understanding of the effects of ozone on mosses as there has been very little information
24 available on this topic in recent years.
9.4.2.2 Summary
25 Visible foliar injury resulting from exposure to O3 has been well characterized and
26 documented over several decades of research on many tree, shrub, herbaceous, and crop
27 species (U.S. EPA. 2006b. 1996b. 1984. 1978a). Ozone-induced visible foliar injury
28 symptoms on certain bioindicator plant species are considered diagnostic as they have
29 been verified experimentally in exposure-response studies, using exposure methodologies
30 such as continuous stirred tank reactors (CSTRs), OTCs, and free-air fumigation.
31 Experimental evidence has clearly established a consistent association of visible injury
32 with O3 exposure, with greater exposure often resulting in greater and more prevalent
33 injury. Since the 2006 O3 AQCD, several multi-year field surveys of O3-induced visible
34 foliar injury have been conducted at National Wildlife Refuges in Maine, Michigan, New
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1 Jersey, and South Carolina. New sensitive species showing visible foliar injury continue
2 to be identified from field surveys and verified in controlled exposure studies.
3 The use of biological indicators in field surveys to detect phytotoxic levels of O3 is a
4 longstanding and effective methodology. The USDA Forest Service through the Forest
5 Health Monitoring (FHM) Program (1990-2001) and currently the Forest Inventory and
6 Analysis (FIA) Program has been collecting data regarding the incidence and severity of
7 visible foliar injury on a variety of O3 sensitive plant species throughout the U.S. The
8 network has provided evidence that O3 concentrations were high enough to induce visible
9 symptoms on sensitive vegetation. From repeated observations and measurements made
10 over a number of years, specific patterns of areas experiencing visible O3 injury
11 symptoms can be identified. As noted in the preceding section, a study of 244 national
12 parks indicated that the risk of visible foliar injury was high in 65 parks (27%), moderate
13 in 46 parks (19%), and low in 131 parks (54%).
14 Evidence is sufficient to conclude that there is a causal relationship between
15 ambient O3 exposure and the occurrence of O3-induced visible foliar injury on
16 sensitive vegetation across the U.S.
9.4.3 Growth, productivity and carbon storage in natural ecosystems
17 Ambient O3 concentrations have long been known to cause decreases in photosynthetic
18 rates, decreases in growth, and decreases in yield (U.S. EPA. 2006b. 1996c. 1986.
19 1978a). The O3-induced damages at the plant scale may translate to the ecosystem scale,
20 and cause changes in productivity and C storage. This section focuses on the responses of
21 C cycling to seasonal or multi-year exposures to O3 from the plant to ecosystem scale.
22 Quantitative responses include changes in plant growth, plant biomass allocation,
23 ecosystem production and ecosystem C sequestration. Because of the available
24 information, most of discussion at the plant scale focuses on the response of individual
25 plants, especially tree seedlings and crops, with limited discussion of mixtures of
26 herbaceous species. Changes at the ecosystem scale are difficult to evaluate directly due
27 to the complexity and the large spatial and temporal scale. The discussion of ecosystem
28 effects focuses on the new studies at the large-scale FACE experiments and on ecological
29 model simulations.
9.4.3.1 Plant growth and biomass allocation
30 The previous O3 AQCDs concluded that there is strong evidence that exposure to O3
31 decreases photosynthesis and growth in numerous plant species (U.S. EPA. 2006b.
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1 1996b. 1984. 1978a). Studies published since the last review support those conclusions
2 and are summarized below.
3 In general, research conducted over several decades has indicated that exposure to O3
4 alters stomatal conductance and reduces photosynthesis in a wide variety of plant species.
5 In a review of more than 55 studies, Wittig et al. (2007) reported that current O3
6 concentrations in the northern hemisphere are decreasing stomatal conductance (13%)
7 and photosynthesis (11%) across tree species. It was also found that younger trees (<4
8 year) were affected less by O3 than older trees. Further, the authors also found that
9 decreases in photosynthesis are consistent with the cumulative uptake of O3 into the leaf.
10 In contrast, several studies reported that O3 exposure may result in loss of stomatal
11 control, incomplete stomatal closure at night and a decoupling of photosynthesis and
12 stomatal conductance, which may have implications for whole- plant water use (Section
13 9.4.5).
14 In a recently published meta-analysis, Wittig et al. (2009) quantitatively compiled peer
15 reviewed studies from the past 40 years on the effect of current and future O3 exposures
16 on the physiology and growth of forest species. Wittig et al. (2009) reported that current
17 ambient O3 concentrations as reported in those studies (-40 ppb) significantly decreased
18 annual total biomass growth (7%) across 263 studies. However, this effect could be
19 greater (11 to 17%) in areas that have higher O3 concentrations and as background O3
20 increases in the future (Wittig et al.. 2009). This meta-analysis demonstrates the
21 coherence of O3 effects across numerous studies and species that used a variety of
22 experimental techniques, and these results support the conclusion of the previous AQCD.
23 In two companion papers, McLaughlin et al. (2007a; 2007b) investigated the effects of
24 ambient O3 on tree growth and hydrology at forest sites in the southern Appalachian
25 Mountains. The authors reported that the cumulative effects of ambient levels of O3
26 decreased seasonal stem growth by 30-50% for most trees species in a high O3 year in
27 comparison to a low O3 year (McLaughlin et al., 2007a). The authors also report that
28 high ambient O3 concentrations can disrupt whole-tree water use and in turn reduce late-
29 season streamflow (McLaughlin et al.. 2007b); see Section 9.4.5 for more on water
30 cycling.
31 Since the 2006 O3 AQCD, several new studies based on the Aspen FACE "free air" O3
32 and CO2 exposure experiment in a forest in Wisconsin were published (Darbah et al..
33 2008: Riikonen et al.. 2008: Darbah et al.. 2007: Kubiske et al.. 2007: Kubiske et al..
34 2006; King et al.. 2005). Over the first seven years of stand development, Kubiske et al.
35 (2006) observed that elevated O3 decreased tree heights, diameters, and main stem
36 volumes in the aspen community by 11, 16, and 20%, respectively. In addition, Kubiske
37 et al. (2007) reported that elevated O3 may change the intra- and inter-species
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1 competition. For example, O3 treatments increased the rate of conversion from a mixed
2 aspen-birch community to a birch dominated community. In a comparison presented in
3 Section 9.6.3 of this document, EPA found that effects on biomass accumulation in aspen
4 during the first seven years closely agreed with the exposure-response function based on
5 data from earlier OTC experiments.
6 Several studies at the Aspen FACE site also considered other growth-related effects of
7 elevated O3. Darbah et al. (2008; 2007) reported that O3 treatments decreased paper birch
8 seed weight and seed germination and that this would likely lead to a negative impact of
9 regeneration for that species. Riikonen et al. (2008) found that elevated O3 decreased the
10 amount of starch in birch buds by 16%, and reduced aspen bud size, which may have
11 been related to the observed delay in spring leaf development. The results suggest that
12 elevated O3 concentrations have the potential to alter C metabolism of overwintering
13 buds, which may have carry-over effects in the subsequent growing season (Riikonen et
14 al.. 2008).
15 Effects on growth of understory vegetation were also investigated at Aspen FACE.
16 Bandeff et al. (2006) found that the effects of elevated CO2 and O3 on understory species
17 composition, total and individual species biomass, N content, and 15N recovery were a
18 result of overstory community responses to those treatments; however, there were no
19 apparent direct treatment effects due to high variability of the data. Total understory
20 biomass increased with increasing light and was greatest under the open canopy of the
21 aspen/maple community, as well as the more open canopy of the elevated O3 treatments
22 (Bandeff et al.. 2006). Similarly, data from a study by Awmack et al. (2007) suggest that
23 elevated CO2 and O3 may have indirect growth effects on red (Trifolium pratense) and
24 white (Trifolium repens) clover in the understory via overstory community effects;
25 however, no direct effects of elevated O3 were observed.
26 Overall, the studies at the Aspen FACE experiment are consistent with many of the OTC
27 studies that were evaluated in previous O3 AQCDs. These results strengthen our
28 understanding of O3 effects on forests and demonstrate the relevance of the knowledge
29 gained from trees grown in open-top chamber studies.
30 For some annual species, particularly crops, the endpoint for an assessment of the risk of
31 O3 exposure is yield or growth, e.g., production of grain. For plants grown in mixtures
32 such as hayfields, and natural or semi-natural grasslands (including native nonagricultural
33 species), endpoints other than production of biomass may be important. Such endpoints
34 include biodiversity or species composition, and effects may result from competitive
35 interactions among plants in mixed-species communities. Most of the available data on
36 non-crop herbaceous species are for grasslands, with many of the recent studies
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1 conducted in Europe. See Section 9.4.7 for a review of the recent literature on O3 effects
2 on competition and biodiversity in grasslands.
Root Growth
3 Although O3 does not penetrate soil, it could alter root development by decreasing
4 C assimilation via photosynthesis leading to less C allocation to the roots (Andersen.
5 2003). The response of root development to O3 exposure depends on available
6 photosynthate within the plant and could vary over time. Many biotic and abiotic factors,
7 such as community dynamics and drought stress, have been found to alter root
8 development under elevated O3. An earlier study at the Aspen FACE experiment found
9 that elevated O3 reduced coarse root and fine roots biomass in young stands of paper
10 birch and trembling aspen (King etal. 2001). However, this reduction disappeared
11 several years later. Ozone significantly increased fine-root (<1.0 mm) in the aspen
12 community (Pregitzer et al., 2008). This increase in fine root production was due to
13 changes in community composition, such as better survival of the O3-tolerant aspen
14 genotype, birch, and maple, rather than changes in C allocation at the individual tree level
15 (Pregitzer et al.. 2008; Zak et al.. 2007). In an adult European beech/Norway spruce
16 forest in Germany, drought was found to nullify the O3-driven stimulation of fine root
17 growth. Ozone stimulated fine-root production of beech during the humid year, but had
18 no significant impact on fine root production in the dry year (Matyssek et al., 2010;
19 Nikolova et al.. 2010).
20 Using a non-destructive method, Vollsnes et al. (2010) studied the in vivo root
21 development of subterranean clover (Trifolium subterraneuni) before, during and after
22 short-term O3 exposure. It was found that O3 reduced root tip formation, root elongation,
23 the total root length, and the ratios between below- and above-ground growth within
24 one week after exposure. Those effects persisted for up to three weeks; however, biomass
25 and biomass ratios were not significantly altered at the harvest five weeks after exposure.
26 Several recent meta-analyses have generally indicated that O3 reduced C allocated to
27 roots. In one meta-analysis, Grantz et al. (2006) estimated the effect of O3 on the
28 root: shoot allometric coefficient (k), the ratio between the relative growth rate of the root
29 and shoot. The results showed that O3 reduced the rootshoot allometric coefficient by
30 5.6%, and the largest decline of the rootshoot allometric coefficient was observed in
31 slow-growing plants. In another meta-analysis including 263 publications, Wittig et al.
32 (2009) found that current O3 exposure had no significant impacts on root biomass and
33 rootshoot ratio when compared to pre-industrial O3 exposure. However, if O3
34 concentrations rose to 81-101 ppb (projected O3 levels in 2100), both root biomass and
35 root:shoot ratio were found to significantly decrease. Gymnosperms and angiosperms
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1 differed in their responses, with gymnosperms being less sensitive to elevated O3. In two
2 other meta-analyses, Wang et al. (2010) found elevated O3 reduced biomass allocation to
3 roots by 8.3% at ambient CO2 and 6.0% at elevated CO2, and Morgan et al. (2003) found
4 O3 reduced root dry weight of soybean. While there is clear evidence that O3 reduced C
5 allocation to roots, results of recent individual studies have been mixed, showing negative
6 (Jones et al.. 2010). non-significant (Andersen et al.. 2010; Phillips et al.. 2009) and
7 positive effects (Pregitzer et al.. 2008; Grebenc and Kraigher. 2007) on root biomass and
8 root: shoot ratio.
9.4.3.2 Summary
9 The previous O3 AQCDs concluded that there is strong and consistent evidence that
10 ambient concentrations of O3 decrease photosynthesis and growth in numerous plant
11 species across the U.S. Studies published since the last review continue to support that
12 conclusion.
13 The meta-analysis by Wittig et al.(2007) and (2009) demonstrates the coherence of O3
14 effects on plant photosynthesis and growth across numerous studies and species using a
15 variety of experimental techniques. Since the 2006 O3 AQCD, several studies were
16 published based on the Aspen FACE experiment using "free air," O3, and CO2 exposures
17 in a forest in Wisconsin. Overall, the studies at the Aspen FACE experiment were
18 consistent with many of the open-top chamber (OTC) studies that were the foundation of
19 previous O3 NAAQS reviews. These results strengthen our understanding of O3 effects
20 on forests and demonstrate the relevance of the knowledge gained from trees grown in
21 open-top chamber studies.
22 In recent studies, O3 was shown to have either negative, non-significant, or positive
23 effects on root biomass and root: shoot ratio. While the findings of individual studies were
24 mixed, recent meta-analyses have generally indicated that O3 reduced C allocated to roots
25 (Wittig et al.. 2009; Grantz et al.. 2006).
26 Evidence is sufficient to conclude that there is a causal relationship between O3
27 exposure and reduced growth of woody and herbaceous vegetation.
9.4.3.3 Reproduction
28 Studies during recent decades have demonstrated O3 effects on various stages of plant
29 reproduction. The impacts of O3 on reproductive development, as reviewed by Black et
30 al. (2000). can occur by influencing (1) age at which flowering occurs, particularly in
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1 long-lived trees that often have long juvenile periods of early growth without flower and
2 seed production; (2) flower bud initiation and development; (3) pollen germination and
3 pollen tube growth; (4) seed, fruit, or cone yields; and (5) seed quality (Table 9-1) (U.S.
4 EPA. 2006b). Several recent studies since the 2006 O3 AQCD further demonstrate the
5 effects of O3 on reproductive processes in herbaceous and woody plant species. Although
6 there have been documented effects of ozone on reproductive processes, a knowledge gap
7 still exists pertaining to the exact mechanism of these responses.
8 Ramo et al. (2007) exposed several meadow species to elevated O3 (40-50 ppb) and CO2
9 (+100 ppm), both individually and combined, over three growing seasons in ground-
10 planted mesocosms, using OTCs. Elevated O3 delayed flowering of Campanula
11 rotundifolia and Vicia cracca. Ozone also reduced the overall number of produced
12 flowers and decreased fresh weight of individual Fragaria vesca berries.
13 Black et al. (2007) exposed Brassica campestris to 70 ppb for two days during late
14 vegetative growth or ten days during most of the vegetative phase. The two-day exposure
15 had no effect on growth or reproductive characteristics, while the 10 day exposure
16 reduced vegetative growth and reproductive site number on the terminal raceme,
17 emphasizing the importance of exposure duration and timing. Mature seed number and
18 weight per pod were unaffected due to reduced seed abortion, suggesting that, although
19 O3 affected reproductive processes, indeterminate species such as B. campestris possess
20 enough compensatory flexibility to avoid reduced seed production (Black et al.. 2007).
21 In the determinate species, Plantago major, Black et al. (2010) found that O3 may have
22 direct effects on reproductive development in populations of differing sensitivity. Only
23 the first flowering spike was exposed to 120 ppb O3 for 7 hours per day on 9 successive
24 days (corresponding to flower development) while the leaves and second spike were
25 exposed to charcoal-filtered air. Exposure of the first spike to O3 affected seed number
26 per capsule on both spikes even though spike two was not exposed. The combined seed
27 weight of spikes one and two was increased by 19% in the two resistant populations,
28 suggesting an overcompensation for injury; whereas, a decrease of 21% was observed in
29 the most sensitive population (Black etal.. 2010). The question remains as to whether
30 these effects are true direct ozone-induced effects or compensatory responses.
31 A study by Darbah et al. (2008; 2007) of paper birch (Betula papyrifera) trees at the
32 Aspen FACE site in Rhinelander, WI investigated the effects of elevated O3 and/or CO2
33 on reproductive fitness. Elevated O3 increased flowering, but decreased seed weight and
34 germination success rate of seeds from the exposed trees. These results suggest that O3
35 can dramatically affect flowering, seed production, and seed quality of paper birch,
36 ultimately affecting its reproductive fitness (Darbah et al., 2008; Darbah et al.. 2007).
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Table 9-1 Ozone effects on plant reproductive processes (derived from Table
AX9-22 of the 2006 ozone AQCD)
Species
Condition Measures
References
Apocynun
androsaemifolium
Flowering time
Bergweiler and Manning (1999)
Buddleia davidii
Flowering time
Findley et al. (1997)
Rubus cuneifolius
Pollen germination
Chappelka (2002)
Plantago major
Pollen tube elongation
Stewart (1998)
Fragaria * ananassa
Fruit yield
Drogoudi and Ashmore (2001): Drogoudi and
Ash mo re (2000)
Plantago major
Seed yield
Lyons and Barnes (1998): Pearson et al. (1996):
Reiling and Davison (1992): Whitfield et al. (1997)
Understory herbs
Seed yield
Harward and Treshow (1975)
Source: Adapted from 2006 O3 AQCD
1
2
3
4
5
6
9
10
11
12
13
14
15
16
17
18
19
20
21
22
9.4.3.4 Ecosystem Productivity and Carbon Sequestration
During the previous NAAQS review, there were limited studies that investigated the
effect of O3 exposure on ecosystem productivity and C sequestration. Recent studies
from long-term FACE experiments provide more evidence of the association of O3
exposure and changes in productivity at the ecosystem scale. In addition to experimental
studies, model studies also assessed the impact of O3 exposure on productivity and
C sequestration from stand to global scales.
Two types of models are most often used to study the ecological consequences of O3
exposure: (1) single plant growth models such as TREGRO and PnET-II (Hogsett et al.,
2008; Martin etal.. 2001; Ollinger et al.. 1997b). and (2) process-based ecosystem
models such as PnET-CN, Dynamic Land Ecosystem Model (DLEM), Terrestrial
Ecosystem Model (TEM), or Met Office Surface Exchange Scheme - Top-down
Representation of Interactive Foliage and Flora Including Dynamics (MOSES-TRIFFID)
(Telzer et al.. 2009; Ren et al.. 2007a: Sitch et al.. 2007; Ollinger et al.. 2002) (Table 9-2).
In these models, carbon uptake is simulated through photosynthesis (TREGRO, PnET -
II, PnET- CN, DLEM and MOSES-TRIFFID) or gross primary production (TEM).
Photosynthesis rate at leaf level is modeled by a function of stomatal conductance and
other parameters in TREGRO, PnET -II, PnET- CN, DLEM and MOSES-TRIFFID.
Photosynthesis at canopy level is calculated by summing either photosynthesis of
different leaf types (TREGRO, DLEM, and MOSES-TRIFFID) or photosynthesis of
different canopy layers (PnET -II, PnET- CN). The detrimental effect of O3 on plant
growth is often simulated by multiplying photosynthesis rate by a coefficient that is
dependent on stomatal conductance and cumulative O3 uptake (Table 9-2). Different
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1 plant functional groups (PFTs, such as deciduous trees, coniferous trees or crops) show
2 different responses to O3 exposure. PnET-II, PnET-CN, TEM, DLEM and MOSES-
3 TRIFFID estimate this difference by modifying net photosynthesis with coefficients that
4 represent the O3 induced fractional reduction of photosynthesis for each functional group.
5 The coefficients used in PnET-II, PnET-CN, TEM, DLEM are derived from the functions
6 of O3 exposure (AOT40) versus photosynthesis reduction from Reich (1987) and
7 Tjoelker et al. (1995). The coefficients used in MOSES-TRIFFID are derived from the
8 O3 dose-photosynthesis response function from Pleijel et al. (2004a) and Karlsson et al.
9 (2004). where O3 dose is estimated by a metric named CUOt (cumulative stomatal uptake
10 of O3). The O3 threshold of CUOt is 1.6 nmol/m2/s for woody PFT and 5 nmol/m2/s for
11 grass PFT, and is different from AOT40, which has an O3 threshold level of 40 ppb for
12 all PFTs. Experimental and model studies on ecosystem productivity and C sequestration
13 at the forest stand scale as well as regional and global scales are reviewed in the
14 following section.
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Table 9-2 Comparison of models used to simulate the ecological
consequences of Os exposure
Model
Model
feature
Carbon uptake
Ozone effect
Reference
TREGRO
Hourly or
daily step,
single plant
model
simulating
vegetation
growth
process
Leaf: leaf photosynthesis is a function of
stomatal conductance, mesophyll
conductance and the gradient of CO2 from
atmosphere to the mesophyll cells
Canopy: Leaf is divided into different ages.
The canopy photosynthesis rate is the sum
the photosynthesis of all foliage groups
The effect of O3 on photosynthesis is Hogsett et al.
simulated by reducing mesophyll (2008):
conductance, and increasing respiration. Weinstein et al.
The degree of O3 damage is determined (2005): Tingey
by ambient O3 exposure, and the et al. (2004)
threshold O3 concentration below which
O3 does not affect mesophyll
conductance and respiration
PnET-ll PnET-ll: Leaf: Maximum photosynthesis rate is
and PnET monthly time- determined by a function of foliar N
-CN step, single concentration, and stomatal conductance is
plant model determined by a function of the actual rate
pnET_CN' of the photosynthesis.
monthly time- Canopy: canopy is divided into multiple,
step, even-mass layers and photosynthesis is
ecosystem simulated by a multilayered canopy
mode submodel
The effect of O3 on photosynthesis is
simulated by an equation of stomatal
conductance and O3 dose (AOT40). The
model assumes that photosynthesis and
stomatal conductance remain coupled
under O3 exposure, with a reduction in
photosynthesis for a given month causing
a proportion reduction in stomatal
conductance.
Ollinger et al.
(2002: 1997b):
Pan et al.
(2009)
TEM monthly time- Ecosystem: TEM is run at a 0.5*0.5 degree
step, resolution. Each grid cell is classified by
ecosystem vegetation type and soil texture, and
mode vegetations and detritus are assumed to
distribute homogeneously within grid cells.
Carbon flows into ecosystem via gross
primary production, which is a function of
maximum rate of assimilation,
photosynthetically active radiation, the leaf
area relative to the maximum annual leaf
area, mean monthly air temperate, and
nitrogen availability.
The direct O3 reduction on GPP is
simulated by multiplying GPP by f(O3)t,
where f(O3)t is determined by
evapotranspiration, mean stomatal
conductance, ambient AOT40, and
empirically O3 response coefficient
derived from previous publications.
Felzer et al.
(2005: 2004)
DLEM daily time- Leaf: photosynthesis is a function of 6
step parameters: photosynthetic photon flux
ecosystem density, stomatal conductance, daytime
model temperature, the atmospheric CO2
concentration, the leaf N content and the
length of daytime.
Canopy: Photosynthetic rates for sunlit leaf
and shaded leaf scale up to the canopy
level by multiplying the estimated leaf area
index
Ecosystem: GPP is the sum of gross C
fixation of different plant function groups
The detrimental effect of O3 is simulated
by multiplying the rate of photosynthesis
by O3eff, where O3eff is a function of
stomatal conductance, ambient AOT40,
and O3 sensitive coefficient. Ozone's
indirect effect on stomatal conductance is
also simulated, with a reduction in
photosynthesis for a given month causing
a reduction in stomatal conductance, and
therefore canopy conductance.
Ren et al.
(2007a:
2007b): Zhang
et al. (2007a)
MOSES- 30 minutes Leaf: photosynthesis is a function of
TRIFFID time-step, environmental and leaf parameters and
dynamic stomatal conductance; Stomatal
global conductance is a function of the
vegetation concentration of CO2 and H2O in air at the
model leaf surface and the current rate of
photosynthesis of the leaf
Canopy: Photosynthetic rates scale up to
the canopy level by multiplying a function of
leaf area index and PAR extinction
coefficient
Ecosystem: GPP is the sum of gross C
fixation of different plant function groups
The effect of O3 is simulated by
multiplying the rate of photosynthesis by
F, where F depends upon stomatal
conductance, O3 exposure, a critical
threshold for O3 damage, and O3
sensitive coefficient (functional type
dependent)
Sitch et al.
(2007)
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Local Scale
1 The above- and below-ground biomass and net primary production (NPP) were measured
2 at the Aspen FACE site after 7 years of O3 exposure. Elevated O3 caused 23, 13 and 14%
3 reductions in total biomass relative to the control in the aspen, aspen-birch and aspen-
4 maple communities, respectively (King et al., 2005). At the Kranzberg Forest FACE
5 experiment in Germany, O3 reduced annual volume growth by 9.5 m3/ha in a mixed
6 mature stand of Norway spruce and European beech (Pretzsch et al., 2010). At the
7 grassland FACE experiment at Alp Flix, Switzerland, O3 reduced the seasonal mean rates
8 of ecosystem respiration and GPP by 8%, but had no significant impacts on aboveground
9 dry matter productivity or growing season net ecosystem production (NEP) (Yolk et al..
10 2011). Ozone also altered C accumulation and turnover in soil, as discussed in Section
11 9.4.6.
12 Changes in forest stand productivity under elevated O3 were assessed by several model
13 studies. TREGRO (Table 9-2) has been widely used to simulate the effects of O3 on the
14 growth of several species in different regions in the U.S. Hogsett et al. (2008) used
15 TREGRO to evaluate the effectiveness of various forms and levels of air quality
16 standards for protecting tree growth in the San Bernardino Mountains of California. They
17 found that O3 exposures at the Crestline site resulted in a mean 20.9% biomass reduction
18 from 1980 to 1985 and 10.3% biomass reduction from 1995 to 2000, compared to the
19 "background" O3 concentrations (O3 concentration in Crook County, Oregon). The
20 level of vegetation protection projected was different depending on the air quality
21 scenarios under consideration. Specifically, when air quality was simulated to just meet
22 the California 8-h average maximum of 70 ppb and the maximum 3 months 12-h SUM06
23 of 25 ppm-h, annual growth reductions were limited to 1% or less, while air quality that
24 just met a previous NAAQS (the second highest 1-h max [125 ppb]) resulted in 6-7%
25 annual reduction in growth, resulting in the least protection relative to background O3
26 (Hogsett etal.. 2008).
27 ZELIG is a forest succession gap model, and has been used to evaluate the dynamics of
28 natural stand succession. Combining TREGRO with ZELIG, Weinstein et al. (2005)
29 simulated the effects of different O3 levels ( 0.5, 1.5, 1.75, and 2 times ambient) on the
30 growth and competitive interactions of white fir and ponderosa pine at three sites in
31 California: Lassen National Park, Yosemite National Park, and Crestline. Their results
32 suggested that O3 had little impact on white fir, but greatly reduced the growth of
33 ponderosa pine. If current O3 concentrations continue over the next century, ambient O3
34 exposure (SUM06 of 110 ppm-h) at Crestline was predicted to decrease individual tree
35 C budget by 10% and decrease ponderosa pine abundance by 16%. Effects at Lassen
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1 National Park and Yosemite National Park sites were found to be smaller because of
2 lower O3 exposure levels (Weinstein et al.. 2005).
3 The effects of O3 on stand productivity and dynamics were also studied by other tree
4 growth or stand models, such as ECOPHYS, INTRASTAND and LINKAGES.
5 ECOPHYS is a functional-structural tree growth model. The model used the linear
6 relationship between the maximum capacity of carboxylation and O3 dose to predict the
7 relative effect of O3 on leaf photosynthesis (Martin et al.. 2001). Simulations with
8 ECOPHYS found that O3 decreased stem dry matter production, stem diameter and leaf
9 dry matter production, induced earlier leaf abscission, and inhibited root growth (Martin
10 et al., 2001). INTRASTAND is an hourly time step model for forest stand carbon and
11 water budgets. LINKAGES is a monthly time step model simulating forest growth and
12 community dynamics. Linking INTRASTAND with LINKAGES, Hanson et al. (2005)
13 found that a simulated increase in O3 concentration in 2100 (a mean 20-ppb increase over
14 the current O3 concentration) yields a 35% loss of net ecosystem C exchange (NEE) with
15 respect to the current conditions (174 g C/m2/year).
Regional and Global Scales
16 Since the publication of the 2006 O3 AQCD, there is additional evidence suggesting that
17 O3 exposure alters ecosystem productivity and biogeochemical cycling at the regional
18 and continental scale. Most of those studies were conducted by using process-based
19 ecosystem models (Table 9-2) and are briefly reviewed in the following sections.
20 Ollinger et al. (1997a) simulated the effect of O3 on hardwood forest productivity of 64
21 hardwood sites in the northeastern U.S. with PnET-II (Table 9-2). Their simulations
22 indicated that O3 caused a 3-16% reduction in NPP from 1987 to 1992 (Table 9-3). The
23 interactive effects of O3, N deposition, elevated CO2 and land use history on C dynamics
24 were estimated by PnET-CN (Table 9-2) (Ollinger etal.. 2002). The results indicated that
25 O3 offset the increase in net C exchange caused by elevated CO2 and N deposition by
26 13% (25.0 g C/m2/year) under agriculture site history, and 23% (33.6 g C/m2/year) under
27 timber harvest site history. PnET-CN was also used to assess changes in C sequestration
28 of U.S. Mid-Atlantic temperate forest. Pan et al. (2009) designed a factorial modeling
29 experiment to separate the effects of changes in atmospheric composition, historical
30 climatic variability and land-disturbances on the C cycle. They found that O3 acted as a
31 negative factor, partially offsetting the growth stimulation caused by elevated CO2 and N
32 deposition in U.S. Mid-Atlantic temperate forest. Ozone decreased NPP of most forest
33 types by 7-8%. Among all the forest types, spruce-fir forest was most resistant to O3
34 damage, and NPP decreased by only 1 % (Pan et al.. 2009).
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1 Felzer et al. (2004) developed TEM 4.3 (Table 9-2) to simulate the effects of O3 on plant
2 growth and estimated effects of O3 on NPP and C sequestration of deciduous trees,
3 conifers and crops in the conterminous U.S. The results indicated that O3 reduced NPP
4 and C sequestration in the U.S. (Table 9-3) with the largest decreases (over 13% in some
5 locations) in NPP occurring in the Midwest agricultural lands during the mid-summer.
6 TEM was also used to evaluate the magnitude of O3 damage at the global scale (Table 9-
7 3) (Felzer etal., 2005). Simulations forthe period 1860 to 1995 show that the largest
8 reductions in NPP and net C exchange occurred in the mid western U.S., eastern Europe,
9 and eastern China (Felzer et al., 2005). DLEM (Table 9-2) was developed to simulate the
10 detrimental effect of O3 on ecosystems, and has been used to examine the O3 damage on
11 NPP and C sequestration in Great Smoky Mountains National Park (Zhang et al., 2007a).
12 grassland ecosystems and terrestrial ecosystems in China (Ren et al.. 2007a: Ren et al..
13 2007b). Results of those simulations are listed in Table 9-3.
14 Instead of using AOT40 as their O3 exposure metric as PnET, TEM and DLEM did,
15 Sitch et al. (2007) incorporated a different O3 metric named CUOt (cumulative stomatal
16 uptake of O3), derived from Pleijel et al. (2004a). into the MOSES-TRIFFID coupled
17 model (Table 9-2). In the CUOt metric, the fractional reduction of plant production is
18 dependent on O3 uptake by stomata over a critical threshold for damage with this
19 threshold level varying by plant functional type. Consistent with previous studies, their
20 model simulation indicated that O3 reduced global gross primary production (GPP),
21 C exchange rate and C sequestration (Table 9-3). The largest reductions in GPP and land-
22 C storage were projected over North America, Europe, China and India. In the model,
23 reduced ecosystem C uptake due to O3 damage, results in additional CO2 accumulation
24 in the atmosphere and an indirect radiative forcing of climate change. Their simulations
25 indicated that the indirect radiative forcing caused by O3 (0.62-1.09 W/m2) could have
26 even greater impact on global warming than the direct radiative forcing of O3
27 (0.89 W/m2) (Sitch et al.. 2007).
28 Results from the various model studies presented in Table 9-3 are difficult to compare
29 because of the various spatial and temporal scales used in these studies. However, all the
30 studies showed that O3 exposure decreased ecosystem productivity and C sequestration.
31 These results are consistent and coherent with experimental results from the leaf, plant
32 and ecosystem scales (Sitch et al., 2007; Felzer etal.. 2005). Many of the models use the
33 same underlying function to simulate the effect of O3 exposure to C uptake. For example
34 the functions of O3 exposure (AOT40) versus photosynthesis reduction for PnET-II,
35 PnET-CN, TEM, DLEM were all from Reich (1987) and Tjoelker et al. (1995).
36 Therefore, it is not surprising that the results are similar. While these models can be
37 improved and more evaluation with experimental data can be done, these models
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1 represent the state of the science for estimating the effect of O3 exposure on productivity
2 and C sequestration.
9.4.3.5 Summary
3 During the previous NAAQS reviews, there were very few studies that investigated the
4 effect of O3 exposure on ecosystem productivity and C sequestration. Recent studies
5 from long-term FACE experiments, such as Aspen FACE, SoyFACE and the Kranzberg
6 Forest (Germany), provided evidence of the association of O3 exposure and reduced
7 productivity at the ecosystem level. Studies at the leaf and plant scales showed that O3
8 reduced photosynthesis and plant growth, which provided coherence and biological
9 plausibility for the decrease in ecosystem productivity. Results across different ecosystem
10 models, such as TREGRO, PnET, TEM and DLEM, were consistent with the FACE
11 experimental evidence, which showed that O3 reduced ecosystem productivity.
12 Although O3 generally causes negative effects on plant growth, the magnitude of the
13 response varies among plant communities. For example, O3 had little impact on white fir,
14 but greatly reduced growth of ponderosa pine in southern California (Weinstein et al..
15 2005). Ozone decreased net primary production (NPP) of most forest types in Mid-
16 Atlantic region, but had small impacts on spruce-fir forest (Pan et al.. 2009).
17 In addition to plant growth, other indicators that are typically estimated by model studies
18 include net ecosystem CO2 exchange (NEE), C sequestration, and crop yield. Model
19 simulations consistently found that O3 exposure caused negative impacts on those
20 indicators, but the severity of these impacts was influenced by multiple interactions of
21 biological and environmental factors. The suppression of ecosystem C sinks results in
22 more CO2 accumulation in the atmosphere. Globally, the indirect radiative forcing caused
23 by O3 exposure through lowering ecosystem C sink could have an even greater impact on
24 global warming than the direct radiative forcing of O3 (Sitch etal.. 2007). Ozone could
25 also affect regional C budgets through interacting with multiple factors, such as N
26 deposition, elevated CO2 and land use history. Model simulations suggested that O3
27 partially offset the growth stimulation caused by elevated CO2 and N deposition in both
28 Northeast- and Mid-Atlantic-region forest ecosystems of the U.S. (Pan et al.. 2009;
29 Ollinger etal.. 2002).
30 The evidence is sufficient to infer that there is a causal relationship between O3
31 exposure and reduced productivity, and a likely causal relationship between O3
32 exposure and reduced carbon sequestration in terrestrial ecosystems.
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Table 9-3 Modeled effects of ozone on primary production, C exchange, and
C sequestration
Scale
GPP Global
Global
U.S.
U.S.
NPP
Northeastern
U.S.
U.S. Mid-
Atlantic
China
Global
C exchange
Global
Global
U.S.
GSM National
Park
C sequestration
China
China
China
Model
MOSES-
TRIFFID
TEM
TEM
TEM
PnET
PnET
DLEM
TEM
MOSES-
TRIFFID
MOSES-
TRIFFID
TEM
DLEM
DLEM
DLEM
DLEM
Index
CUOta
AOT40
AOT40
AOT40
AOT40
AOT40
AOT40
AOT40
cuot
cuot
AOT40
AOT40
AOT40
AOT40
AOT40
Ozone Impacts
Decreased by 14-23% over the period 1901-2100
Decreased by 0.8% without agricultural
management and a decrease of 2.9% with optimal
agricultural management
Reduced by 2.3% without optimal N fertilization and
7.2% with optimal N fertilization from 1983-1993
Reduced by 2.6-6.8% during the late 1980s-early
1990s.
A reduction of 3-1 6% from 1 987-1 992
Decreased NPP of most forest types by 7-8%
Reduced NPP of grassland in China by 8.5 Tg C
from 1960s to 1990s
Reduced net C exchange (1950-1995) by 0.1 Pg
C/yr without agricultural management and 0.3 Pg
C/yr with optimal agricultural management
Decreased global mean land-atmosphere C fluxes
by 1 .3 Pg C/yr and 1 .7 Pg C/yr for the 'high' and
'low' plant O3 sensitivity models, respectively
Reduced land-C storage accumulation by between
143 Pg C/yr and 263 Pg C/yr from 1900-2100
Reduced C sequestration by 1 8-38 Tg C/yr from
1950to1995
Decreased the ecosystem C storage of deciduous
forests by 2.5% and pine forest by 1 .4% from 1 971
to 2001
Reduced total C storage by 0.06% in 1960s and
1 .6% in 1990s in China's terrestrial ecosystems
O3 exposure reduced the net C sink of China's
terrestrial ecosystem by 7% from 1961 to 2005
Ozone induced net carbon exchange reduction
ranged from 0.4-43.1% , depending on different
forest type
Reference
Sitch et al.
(2007)
Felzer et al.
(2005)
Felzer et al.
(2005)
Felzer et al.
(2004)
Ollinger et al.
(1997a)
Pan et al.
(2009)
Ren et al.
(2007b)
Felzer et al.
(2005)
Sitch et al.
(2007)
Sitch et al.
(2007)
Felzer et al.
(2004)
Zhang et al.
(2007a)
Ren et al.
(2007a)
Tian et al.
(2011)
Ren et al.
(2011)
aCUOt is defined as the cumulative stomatal uptake of Os, using a constant Os-uptake rate threshold oft nmol/m /s.
dPg equals 1 x 1015 grams.
2
O
4
5
6
9.4.4 Crop yield and quality in agricultural systems
The detrimental effect of O3 on crop production has been recognized since the 1960s and
a large body of research has stemmed from that recognition. Previous O3 AQCDs have
extensively reviewed this body of literature. Table 9-4 summarizes recent experimental
studies of O3 effects on agricultural crops, exclusive of growth and yield. Growth and
yield results are summarized in Table 9-17.
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September 2011
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1 The actual concentration and duration threshold for O3 damage varies from species to
2 species and sometimes even among genotypes of the same species (Guidi et al.. 2009;
3 Sawada and Kohno. 2009; Biswas et al.. 2008; Arivaphanphitak et al.. 2005; Dalstein and
4 Vas. 2005; Keutgen et al.. 2005). A number of comprehensive reviews and meta-analyses
5 have recently been published discussing both the current understanding of the
6 quantitative effects of O3 concentration on a variety of crop species and the potential
7 focus areas for biotechnological improvement to a future growing environment that will
8 include higher O3 concentrations (Bender and Weigel. 2011; Booker et al.. 2009;
9 Van Dingenen et al.. 2009; Ainsworth. 2008; Feng et al.. 2008b; Haves et al.. 2007; Mills
10 et al.. 2007b; Grantz et al.. 2006; Morgan etal.. 2003). Since the 2006 O3 AOCDOJ.S.
11 EPA. 2006b). exposure-response indices for a variety of crops have been suggested
12 (Mills et al.. 2007b) and many reports have investigated the effects of O3 concentration
13 on seed or fruit quality to extend the knowledge base beyond yield quantity. This section
14 will outline the key findings from these papers as well as highlight some of the recent
15 research addressing the endpoints such as yields and crop quality.
16 This section will also highlight recent literature that focuses on O3 damage to crops as
17 influenced by other environmental factors. Genetic variability is not the only factor that
18 determines crop response to O3 damage. Ozone concentration throughout a growing-
19 season is not homogeneous and other environmental conditions such as elevated CO2
20 concentrations, drought, cold or nutrient availability may alleviate or exacerbate the
21 oxidative stress response to a given O3 concentration.
9.4.4.1 Yield
22 It is well known that yield is negatively impacted in many crop species in response to
23 high O3 concentration. However, the concentrations at which damage is observed vary
24 from species to species. Numerous analyses of experiments conducted in OTCs and with
25 naturally occurring gradients demonstrate that the effects of O3 exposure also vary
26 depending on the growth stage of the plant; plants grown for seed or grain are often most
27 sensitive to exposure during the seed or grain-filling period (Sojaetal.. 2000; Pleijel et
28 al.. 1998; Younglove et al.. 1994; Leeetal.. 1988a). AX9.5.4.1 of the 2006 O3 AQCD
29 summarized many previous studies on crop yield.
Field studies and meta-analyses
30 The effect of O3 exposure on U.S. crops remains an important area of research and
31 several studies have been published on this topic since the 2006 O3 AQCD (U.S. EPA.
32 2006b) (Table 9-4 and 9-17). For example, one study with cotton in a crop-weed
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1 interaction study (Grantz and Shrestha. 2006) utilizing OTCs suggests that 12-hour
2 average O3 concentrations of 79.9 ppb decreased cotton biomass by 25% and 12-hour
3 average O3 concentration of 122.7 ppb decreased cotton biomass by 75% compared to
4 charcoal filtered control (12-h avg: 12.8 ppb). Further, this study suggests that the weed,
5 yellow nutsedge, was less sensitive to increasing O3 concentration, which would increase
6 weed competition (Grantz and Shrestha. 2006). In a study of peanuts in North Carolina,
7 near ambient and elevated exposures of O3 reduced photosynthesis and yield compared to
8 very low O3 conditions (Booker et al.. 2007; Burkey et al.. 2007). In another study,
9 Grantz and Vu (2009) reported that sugarcane biomass growth significantly declined
10 under O3 exposure.
11 The average yield loss reported across a number of meta-analytic studies have been
12 published recently for soybean (Morgan et al.. 2003). wheat (Feng et al.. 2008b). rice
13 (Ainsworth. 2008). semi-natural vegetation (Hayes et al.. 2007). potato, bean and barley
14 (Feng and Kobayashi. 2009). Meta-analysis allows for the objective development of a
15 quantitative consensus of the effects of a treatment across a wide body of literature.
16 Further, this technique allows for a compilation of data across a range of O3 fumigation
17 techniques, durations and concentrations in order to assemble the existing literature in a
18 meaningful manner.
19 Morgan et al. (2003) reported an average seed yield loss for soybean of 24% compared to
20 charcoal filtered air across all O3 concentrations used in the 53 compiled studies. The
21 decrease in seed yield appeared to be the product of nearly equal decreases (7-12%) in
22 seed weight, seed number and pod number. As would be expected, the lowest O3
23 concentration (30-59 ppb) resulted in the smallest yield losses, approximately 8%, while
24 the highest O3 concentration (80-120 ppb ) resulted in the largest yield losses,
25 approximately 35% (Morgan et al.. 2003). Further, the oil/protein ratio within the
26 soybean seed was altered due to growth at elevated O3 concentrations, with a decrease in
27 oil content. The studies included in this meta-analysis all used enclosed fumigation
28 systems or growth chambers which may have altered the coupling of the atmosphere to
29 the lower plant canopy (McLeod and Long. 1999), although the results of Morgan et al.
30 (2006). Betzelberger et al. (2010). and the comparisons presented in Section 9.6.3
31 strongly suggest that decreases in yield between ambient and elevated exposures are not
32 affected by exposure method. Utilizing the Soybean Free Air gas Concentration
33 Enrichment Facility (SoyFACE; www.soyface.illinois.edu). Morgan et al. (2006) report a
34 20% seed yield loss due to a 23% increase in average daytime O3 concentration
35 (56-69 ppb) within a single soybean cultivar across two growing seasons in Illinois,
36 which lies within the range predicted by the meta-analysis. A further breakdown of the
37 effects of current O3 concentrations (AOT40 of 4.7 ppm-h) on bean seed quality
38 (Phaseolus vulgaris) has identified that growth at current O3 concentrations compared to
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1 charcoal-filtered air raised total lipids, total crude protein and dietary fiber content (Iriti et
2 al.. 2009). An increase in total phenolics was also observed, however the individual
3 phenolics compounds responded differently, with significant decreases in anthocyanin
4 content. The seeds from ambient O3 exposed plants also displayed increased total
5 antioxidant capacity compared to charcoal-filtered air controls (Iriti et al.. 2009).
6 Betzelberger et al. (2010) has recently utilized the SoyFACE facility to compare the
7 impact of elevated O3 concentrations across 10 soybean cultivars to investigate
8 intraspecific variability of the O3 response to find physiological or biochemical markers
9 for eventual O3 tolerance breeding efforts (Betzelberger et al.. 2010). They report an
10 average 17% decrease in yield across all 10 cultivars across two growing seasons due to a
11 doubling of ambient O3 concentrations, with the individual cultivar responses ranging
12 from -7% to -36%. The exposure-response functions derived for these 10 current
13 cultivars were similar to the response functions derived from the NCLAN studies
14 conducted in the 1980s (Heagle. 1989) suggesting there has not been any selection for
15 increased tolerance to O3 in more recent cultivars. More complete comparisons between
16 yield predictions based on data from cultivars used in NCLAN studies, and yield data for
17 modern cultivars from SoyFACE are reported in Section 9.6.3 of this document. They
18 confirm that the response of soybean yield to O3 exposure has not changed in current
19 cultivars.
20 A meta-analysis has also been performed on studies investigating the effects of O3
21 concentrations on wheat (Feng et al., 2008b). Across 23 studies included, elevated O3
22 concentrations (ranging from a 7-h daily average of 31-200 ppb) decreased grain yield by
23 29%. Winter wheat and spring wheat did not differ in their responses; however the
24 response in both varieties to increasing O3 concentrations resulted in successively larger
25 decreases in yield, from a 20% decrease in 42 ppb to 60% in 153 ppb O3. These yield
26 losses were mainly caused by a combination of decreases in individual grain weight (-
27 18%), ear number per plant (-16%), and grain number per ear (-11%). Further, the grain
28 starch concentration decreased by 8% and the grain protein yield decreased by 18% due
29 to growth at elevated O3 concentrations as well. However, increases in grain calcium and
30 potassium levels were reported (Fenget al.. 2008b).
31 A recent meta-analysis found that growth at elevated O3 concentrations negatively
32 impacts nearly every aspect of rice performance as well (Ainsworth. 2008). While rice is
33 not a major crop in the U.S., it provides a staple food for over half of the global
34 population (IRRI. 2002) and the effects of rising O3 concentrations on rice yields merits
35 consideration. On average, rice yields decreased 14% in 62 ppb O3 compared to charcoal-
36 filtered air. This yield loss was largely driven by a 20% decrease in grain number
37 (Ainsworth. 2008).
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1 Feng and Kobayashi (2009) have recently compiled yield data for six major crop species,
2 potato, barley, wheat, rice, bean and soybean and grouped the O3 treatments used in those
3 studies into three categories: baseline O3 concentrations (<26 ppb), current ambient 7- or
4 12-h daily O3 concentrations (31-50 ppb), and future ambient 7- or 12-h daily O3
5 concentrations (51-75 ppb). Using these categories, they have effectively characterized
6 the effects of current O3 concentrations and the effects of future O3 concentrations
7 compared to baseline O3 concentrations. At current O3 concentrations, which ranged
8 from 41-49 ppb in the studies included, soybean (-7.7%), bean (-19.0%), barley (-8.9%),
9 wheat (-9.7%), rice (-17.5%) and potato (-5.3%) all had yield losses compared to the
10 baseline O3 concentrations (<26 ppb). At future O3 concentrations, averaging 63 ppb,
11 soybean (-21.6%), bean (-41.4%), barley (-14%), wheat (-28%), rice (-17.5%) and potato
12 (-11.9%) all had significantly larger yield losses compared to the losses at current O3
13 concentrations (<26 ppb) (Feng and Kobayashi, 2009).
14 A review of OTC studies has determined the AOT40 critical level that causes a 5% yield
15 reduction across a variety of agricultural and horticultural species (Mills et al.. 2007b).
16 The authors classify the species studied into three groups: sensitive, moderate and
17 tolerant. The sensitive crops, including watermelon, beans, cotton, wheat, turnip, onion,
18 soybean, lettuce, and tomato, respond with a 5% reduction in yield under a 3-month
19 AOT40 of 6 ppm-h. Watermelon was the most sensitive with a critical level of
20 1.6 ppm-h. The moderately sensitive crops, including sugar beet, oilseed rape, potato,
21 tobacco, rice, maize, grape and broccoli, responded with a 5% reduction in yield between
22 8.6 and 20 ppm-h. The crops classified as tolerant, including strawberry, plum and barley,
23 responded with a 5% yield reduction between 62-83.3 ppm-h (Mills et al.. 2007b).
24 Feng and Kobayashi (2009) compared their exposure-response results to those published
25 by Mills et al. (2007b) and found the ranges of yield loss to be similar for soybean, rice
26 and bean. However, Feng and Kobayasi (2009) reported smaller yield losses for potato
27 and wheat and larger yield losses for barley compared to the dose-response functions
28 published by Mills et al. (2007b), which they attributed to their more lenient criteria for
29 literature inclusion.
30 While the studies investigating the impact of various O3 concentrations on yield are
31 important and aid in determining the vulnerability of various crops to a variety of O3
32 concentrations, there is still uncertainty as to how these crops respond under field
33 conditions with interacting environmental factors such as temperature, soil moisture, CO2
34 concentration, and soil fertility (Booker et al.. 2009). Further, there appears to be a
35 distinct developmental and genotype dependent influence on plant sensitivity to O3 that
36 has yet to be fully investigated across O3 concentrations in a field setting. The potentially
37 mitigating effect of breeding selection for O3 resistance has received very little attention
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1 in the published scientific literature. Anecdotal reports suggest that such selection may
2 have occurred in recent decades for some crops in areas of the country with high ambient
3 exposures. However, the only published literature available is on soybean and these
4 studies indicate that sensitivity has not changed in cultivars of soybean between the
5 1980s and the 2000s (Betzelberger et al.. 2010). This conclusion for soybeans is
6 confirmed by comparisons presented in Section 9.6.3 of this document.
Yield loss at regional and global scales
7 Because O3 is heterogeneous in both time and space and O3 monitoring stations are
8 predominantly near urban areas, the impacts of O3 on current crop yields at large spatial
9 scales are difficult to estimate. Fishman et al. (2010) have used satellite observations to
10 estimate O3 concentrations in the contiguous tri-state area of Iowa, Illinois and Indiana
11 and have combined that information with other measured environmental variables to
12 model the historical impact of O3 concentrations on soybean yield across the 2002-2006
13 growing seasons. When soybean yield across Iowa, Indiana and Illinois was modeled as a
14 function of seasonal temperature, soil moisture and O3 concentrations, O3 had the largest
15 contribution to the variability in yield for the southern-most latitudes included in the
16 dataset. Fishman et al. (2010) determined that O3 concentrations significantly reduced
17 soybean yield by 0.38 to 1.63% for every additional ppb of exposure across the 5 years.
18 This value is consistent with previous chamber studies (Heagle. 1989) and results from
19 SoyFACE (Morgan et al.. 2006). Satellite estimates of tropospheric O3 concentrations
20 exist globally (Fishman et al., 2008). therefore utilizing this historical modeling approach
21 is feasible across a wider geographical area, longer time-span and perhaps for more crop
22 species.
23 The detrimental effects of O3 on crop production at regional or global scales were also
24 assessed by several model studies. Two large scale field studies were conducted in the
25 U.S. (NCLAN) and in Europe (European Open Top Chamber Programme, EOTCP) to
26 assess the impact of O3 on crop production. Ozone exposure-response regression models
27 derived from the two programs have been widely used to estimate crop yield loss
28 (Avnerv et al.. 201 la. b; Van Dingenen et al.. 2009: Tong and Mauzerall. 2008: Wang
29 and Mauzerall. 2004). Those studies found that O3 generally reduced crop yield and that
30 different crops showed different sensitivity to O3 pollution (Table 9-5). Ozone was
31 calculated to induce a possible 45-82 million metric tons loss for wheat globally.
32 Production losses for rice, maize and soybean were on the order of 17-23 million metric
33 tons globally (Van Dingenen et al.. 2009). The largest yield losses occur in high-
34 production areas exposed to high O3 concentrations, such the Midwest and the
35 Mississippi Valley regions in the U.S., Europe, China and India (Van Dingenen et al..
36 2009: Tong etal.. 2007).
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9.4.4.2 Crop Quality
1 In general, it appears that increasing O3 concentrations above current ambient
2 concentrations can cause species-dependent biomass losses, decreases in root biomass
3 and nutritive quality, accelerated senescence and shifts in biodiversity. A study conducted
4 with highbush blackberry has demonstrated decreased nutritive quality with increasing
5 O3 concentration despite no change in biomass between charcoal-filtered control,
6 ambient O3 and 2 x ambient O3 exposures (Ditchkoff et al.. 2009). A study conducted
7 with sedge using control (30 ppb), low (55 ppb), medium (80 ppb) and high (105 ppb) O3
8 treatments has demonstrated decreased root biomass and accelerated senescence in the
9 medium and high O3 treatments (Jones et al.. 2010). Alfalfa showed no biomass changes
10 across two years of double ambient O3 concentrations (AOT40 of 13.9 ppm-h) using
11 FACE fumigation (Maggio et al.. 2009). However a modeling study has demonstrated
12 that 84% of the variability in the relative feed value in high-yielding alfalfa was due to
13 the variability in mean O3 concentration from 1998-2002 (Lin et al.. 2007). Further, in a
14 managed grassland FACE system, the reduction in total biomass harvest over five years
15 decreased twice as fast in the elevated treatment (AOT40 of 13-59 ppm-h) compared to
16 ambient (AOT40 of 1-20.7 ppm-h). Compared with the ambient control, loss in annual
17 dry matter yield was 23% after 5 year. Further, functional groups were differentially
18 affected, with legumes showing the strongest negative response (Volketal.. 2006).
19 However, a later study by Stampfli and Fuhrer (2010) at the same site suggested that
20 Volk et al.(2006) was likely overestimated the effects of O3 on yield reduction because
21 the overlapping effects of species dynamics caused by heterogeneous initial conditions
22 and a change in management were not considered in Volk et al. (2006). An OTC study
23 conducted with Trifolium subterraneum exposed to filtered ( <15 ppb), ambient, and
24 40 ppb above ambient O3 demonstrates decreases in biomass in the highest O3 treatment
25 as well as 10-20% decreased nutritive quality which was mainly attributed to accelerated
26 senescence (Sanz et al.. 2005). A study conducted with Eastern gamagrass and big
27 bluestem in OTCs suggested that big bluestem is not sensitive to O3, but gamagrass
28 displayed decreased nutritive quality in the 2 x ambient O3 treatment, due to higher
29 lignin content and decreased N, (Lewis etal. 2006).
9.4.4.3 Summary
30 The detrimental effect of O3 on crop production has been recognized since the 1960's
31 and a large body of research has subsequently stemmed from those initial findings.
32 Previous O3 AQCDs have extensively reviewed this body of literature (U.S. EPA,
33 2006b). Current O3 concentrations across the U.S. are high enough to cause yield loss for
34 a variety of agricultural crops including, but not limited to, soybean, wheat, potato,
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1 watermelon, beans, turnip, onion, lettuce, and tomato. Continued increases in O3
2 concentration may further decrease yield in these sensitive crops. Despite the well-
3 documented yield losses due to increasing O3 concentration, there is still a knowledge
4 gap pertaining to the exact mechanisms of O3-induced yield loss. Research has linked
5 increasing O3 concentration to decreased photosynthetic rates and accelerated
6 senescence, which are related to yield.
7 New research is beginning to consider the mechanism of damage caused by prolonged,
8 lower O3 concentration (so-called chronic exposure) compared to short, very high O3
9 concentration (so-called acute exposure). Both types of O3 exposure cause damage to
10 agricultural crops, but through very different mechanisms. Historically, most research on
11 the mechanism of O3 damage used acute exposure studies. During the last decade, it has
12 become clear that the cellular and biochemical processes involved in the response to
13 acute O3 exposure are not involved in response to chronic O3 exposure, even though both
14 cause yield loss in agriculturally important crops.
15 In addition, new research has highlighted the effects of O3 on crop quality. Increasing O3
16 concentration decreases nutritive quality of grasses, decreases macro- and micro-nutrient
17 concentrations in fruits and vegetable crops, and decreases cotton fiber quality. These
18 areas of research require further investigation to determine mechanisms and exposure-
19 response relationships.
20 During the previous NAAQS reviews, there were very few studies that estimated O3
21 impacts on crop yields at large spatial scales. Recent modeling studies found that O3
22 generally reduced crop yield, but the impacts varied across regions and crop species. For
23 example, the largest O3-induced crop yield losses occurred in high-production areas
24 exposed to high O3 concentrations, such the Midwest and the Mississippi Valley regions
25 of the U.S. (Van Dingenen et al.. 2009). Among crop species, the estimated yield loss for
26 wheat and soybean were higher than for rice and maize (Van Dingenen et al.. 2009).
27 Using satellite air-column observations with direct air-sampling O3 data, Fishman et al.
28 (2010) modeled the yield-loss due to O3 over the continuous tri-state area of Illinois,
29 Iowa and Wisconsin. They determined that O3 concentrations significantly reduced
30 soybean yield, which further reinforces previous results from FACE-type experiments
31 and OTC experiments.
32 Evidence is sufficient to conclude that there is a causal relationship between O3
33 exposure and reduced yield and quality of agricultural crops.
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Table 9-4
Species
Facility
Location
Alfalfa (Medicago
sativa cv. Beaver)
Growth chambers
Bean
(Phaseolus vulgaris \.
cv Borlotto)
OTC, ground-planted
Curno, Italy
Big Blue Stem
(Andropogon gerardif)
OTC
Alabama, U.S.
Brassica napus
Growth chambers
Belgium
Brassica napus
cv. Westar
Growth chambers
Finland
Eastern Gamagrass
(Tripsacum
dactyloides)
OTC
Alabama, U.S.
Lettuce
(Lactuca sativa)
OTC
Carcaixent
Experimental Station,
Spain
Peanut
(Arachis hypogaea)
OTC
Raleigh, NC; U.S.
Poa pratensis
OTC
Braunschweig,
Germany
Potato
(Solarium tuberosum
cv. Bintje
OTC
Sweden & Finland
Potato
(Solanum tuberosum
cv. Indira)
Climate chambers
Germany
Soybean
OTC
Italy
Summary of recent studies of ozone effects on crops (exclusive of
growth and yield)
Exposure
Duration
1,2 or
4 days
4 months
4 months
4 days
17-26 days
4 months
30 days
Syr
3yr;
4-5 wk
in the
spring
2yr
8wk
Syr
Ozone Exposure9
(Additional treatment)
3, 5 or - h/day
85ppb
(Exposure duration)
Seasonal AOT40:
CF = 0.5ppm-h;
Ambient = 4.6 ppm-h
(N/A)
12-havg:
CF=14ppb;
Ambient = 29 ppb;
Elevated = 71 ppb
(N/A)
CF&176ppb
for 4 h/day
(N/A)
8-h avg:
CF&100ppb
(Bt/non-Bt;
herbivory)
12-havg:
CF=14ppb;
Ambient = 29 ppb;
Elevated = 71 ppb
(N/A)
12-h mean:
CF= 10.2 ppb;
NF = 30.1 ppb;
NF+03 = 62.7 ppb
(4 cultivars)
12-havg:
CF = 22 ppb;
Ambient = 46 ppb;
Elevated = 75 ppb
(C02:375ppm;
548 ppm; 730 ppm)
8-h avg:
CF+25=21.7ppb;
NF+50=73.1 ppb
(Competition)
CF=10ppb;
Ambient = 25 ppb);
Ambient(+) = (36 ppb);
Ambient(++) = (47 ppb)
(N/A)
CF=10ppb;
Ambient = 50 ppb;
2xAmbient= 100 ppb
(CO 2 : 400 ppm &
700 ppm)
AOT40:
CF = 0 ppm-h;
Ambient = 3.4 ppm-h;
Elevated = 9.0 ppm-h
(Well-watered &
water-stressed)
Variable(s) measured
Relative feed value
Seed lipid,
Protein content
Fiber content
Relative feed value
Glucosinolates
VOC emissions
Relative feed value
Lipid peroxidation;
Root length
Harvest biomass
Relative feed value
[K],[Ca],[Mg],[P],[N]perdry
weight of tubers *dose-response
regression, report significant
positive or negative slope with
increasing [03]
Pathogen infestation using %
necrosis
Daily
evapotranspiration
percent change from
(percent change from
ambient)
n.s.
"high variability among
treatment groups (N/A)
+28.5 (N/A)
+7.88 (N/A)
+14.54 (N/A)
n.s. (n.s.)
-41 (N/A)
-30.7 (N/A);
-34 (N/A)
-17 (-12)
+77 (+38)
-22 (-14)
-40 (-10)
N/A (n.s.; -8)
[N] [P] [Ca] n.s.;
[K]&[Mg]sig +
(N/A)
+52 (n.s.)
-28 (-14)
Reference
Muntifering etal.
(2006)
Iriti etal. (2009)
Lewis etal. (2006)
Gielen etal. (2006)
Himanen et al.
(2009b)
Lewis et al. (2006)
Calatayud et. al.
(2002)
Booker etal.
(2007)
Bender etal.
(2006)
Piikki et al. (2007)
Plessl etal. (2007)
Jaude et al. (2008)
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Species
Facility
Location
Soybean
(Glycine max
cv. 93B15)
SoyFACE
Urbana, IL; U.S.
Soybean
(Glycine max
CV.93B15)
SoyFACE
Urbana, IL; U.S.
Soybean
(Glycine max
cv. Essex)
OTC, ground-planted
Raleigh, NC; U.S.
Soybean
(Glycine max
cv. Essex)
OTCs, 21 L pots
Raleigh, NC; U.S.
Soybean
(Glycine max)
lOcultivars)
SoyFACE
Urbana, IL; U.S.
Spring Wheat
(Triticum aestivum
cv. Minaret; Satu;
Drabant; Dragon)
OTCs
Belgium, Finland,
& Sweden
Strawberry
(Fragaria x ananassa
Duch. Cv. Korona
& Elsanta)
Growth chambers
Bonn, Germany
Sweet Potato
Growth Chambers
Bonn, Germany
Tomato
(Lycopersicon
esculentum)
OTC
Valencia, Spain
Trifolium repens &
Trifolium pretense
Aspen FACE
Rhinelander,WI;U.S.
Exposure
Duration
Syr
May-Oct
4 months
2yr
2x3
months
2yr
7yr
2 months
4wk
133 days
3 months
Ozone Exposure9
(Additional treatment)
AOT40:
Ambient = 5-22 ppm-h;
Elevated = 20-43 ppm-h
(C02:550ppm;
environmental
variability)
8-h avg:
Ambient = 38.5 ppb;
Elevated = 52 ppb
(Herbivory)
12-havg:
CF = 21 ppb;
1.5xAmbient = 74 ppb
(C02:370ppm&
714ppm)
12-havg:
CF=18ppb);
elevated - 11 ppb)
(CO 2 : 367 & 71 8)
8-h avg (ppb):
Ambient = 46.3 & 37.9;
Elevated = 82.5 & 61 .3
(Cultivar comparisons)
Seasonal AOT40s
ranged from
Oto16ppm-h
(N/A)
8-h avg:
CF = 0 ppb;
Elevated = 78 ppb
(N/A)
8-h avg:
CF = 0 ppb;
Ambient < 40 ppb;
Elevated = 255 ppb
(N/A)
8- mean:
CF= 16.3 ppb;
NF = 30.1 ppb;
NF(+) = 83.2 ppb
(Various cultivars;
early & late harvest)
3-mo daylight avg:
Ambient = 34.8 ppb;
1.2xAmbient = 42.23
ppb
(C02;560ppm)
Variable(s) measured
Photosynthesis in new leaves,
Herbivory
defense-related
genes
Post-harvest residue
Water-use efficiency
Total antioxidant capacity
Seed protein content;
1 ,000-seed weight regressed
across all experiments
Total leaf area
Tuber weight
Brix degree
Lignin;
Dry-matter
digestibility
percent change from
(percent change from
ambient)
N/A (n.s.)
N/A (N/A)
N/A (-15.46)
n.s. (N/A)
N/A (+19)
N/A (Significant negative
correlation)
N/A (Significant negative
correlation)
-16 (N/A)
-14 (-11. 5)
-7.2 (-3.6)
N/A (+19.3)
N/A (-4.2)
Reference
Bernacchietal.
(2006)
Casteel et al.
(2008)
Booker etal.
(2005)
Booker etal.
(Booker etal..
2004a)
Betzelberger et al.
(2010)
Piikkietal. (2008a)
Keutgen et al.
(2005)
Keutgen et al.
(2008)
Calvo, et al. (2005)
Muntifering et al.
(2006)
aOzone exposure in ppb unless otherwise noted.
bCF = Carbon-filtered air.
NF = Non-filtered air.
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Table 9-5 Modeled effects of ozone on crop yield loss at regional and global
scales
Scale
Global
Global
U.S.
U.S.
East
Asia
Index
M7a;M12b;
AOT40
M12b;AOT40
M7;M12;AOT40
SUM06
M7;M12
Ozone Impacts
Reduced by 7.3% to 12.3% for wheat, 5.4% to 15.6% for soybean, 2.8% to 3.7% for rice,
and 2.4% to 4.1% for maize in year 2000.
Ovinduced global yield reductions ranged from 8.5-1 4% for soybean, 3.9-15% for wheat,
and 2.2-5.5% for maize in year 2000. Global crop production losses totaled 79-121 million
metric tons, worth $11-18 billion annually (USD2000).
Reduced by 4.1 % to 4.4% for wheat, 7.1 % to 1 7.7% for soybean, 2.6% to 3.2% for rice,
and 2.2% to 3.6% for maize in year 2000.
Caused a loss of 53.8 million to 438 million bushels in soybean production, which account
for 1 .7-14.2% of total U.S. soybean production in 2005
Reduced the yield of wheat, rice and corn by 1-9% and soybean by 23-27% in China,
Japan and South Korea in 1990
Reference
Van Dingenenetal.
(2009)
Avneryetal. (2011 a)
Van Dingenen et al.
(2009)
long et al. (2007)
Wang and Mauzerall
(2004)
aM7 is defined as 7-h mean 03 concentration (ppb).
bM12 is defined as 12-h mean 03 concentration (ppb).
1
2
3
4
5
6
7
9.4.5 Water Cycling
Ozone can affect water use in plants and ecosystems through several mechanisms
including damage to stomatal functioning and loss of leaf area. Section 9.3.2 reviewed
possible mechanisms for effects of O3 exposure on stomatal functioning including build-
up of CO 2 in substomatal cavity, impacts on signal transduction pathways, and direct O3
impact on guard cells. Regardless of the mechanism, O3 exposure has been shown to alter
stomatal performance, which may affect plant and stand transpiration and therefore could
affect hydrological cycling (Figure 9-7).
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O3 exposure
Decrease stomatal
conductance or
sluggish stomatal
^response
Altered canopy
water loss
Figure 9-7 The potential effects of ozone exposure on watering cycling.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
In the literature, there is not a clear consensus on the nature of leaf-level stomatal
conductance response to O3 exposure. At the leaf level, O3 exposure is known to result in
stomatal patchiness (Paoletti and Grulke. 2005; Omasa etal.. 1987; Ellenson and
Amundson. 1982). i.e., the heterogeneous aperture of stomata on the leaf surface, and, as
a result, the collective response of groups of stomata on leaves and canopies determines
larger-scale responses to O3. When measured at steady-state high light conditions, leaf-
level stomatal conductance is often found to be reduced when exposed to O3. For
example, a meta-analysis of 55 studies found that O3 reduced stomatal conductance by
11% (Wittig etal.. 2007). However, these steady-state measurements were generally
taken at saturating light conditions and steady-state vapor pressure deficit (VPD).
Saturating light and steady-state VPD conditions are not common in the field since many
parts of the plant canopy are shaded throughout the day. When studied under varying
environmental conditions, many studies have reported incomplete stomatal closure with
elevated O3 exposure during the day (Mills et al.. 2009; Grulke et al.. 2007b; Matyssek et
al.. 1995; Wieser and Havranek. 1995) or at night (Grulke et al.. 2004). This may be due
to sluggish stomatal response. Sluggish stomatal response, defined as a delay in stomatal
response to changing environmental factors relative to controls (Paoletti and Grulke.
2010) has also been documented by several researchers (Grulke et al.. 2007c; Matyssek et
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1 al.. 1995; Pearson and Mansfield. 1993; Wallin and Skarbv. 1992; Lee et al.. 1990;
2 Skarbyetal.. 1987; Keller and Hasler. 1984; Reich and Lassoie. 1984). Sluggish stomatal
3 response associated with O3 exposure suggests an uncoupling of the normally tight
4 relationship between carbon assimilation and stomatal conductance as measured under
5 steady-state conditions (Gregg et al.. 2006; Paoletti and Grulke. 2005). Several tree and
6 ecosystem models, such as TREGRO, PnET and DLEM, rely on this tight relationship to
7 simulate water and carbon dynamics. The O3-induced impairment of stomatal control
8 may be more pronounced for plants growing under water stress (Wilkinson and Davies.
9 2010; Grulke et al.. 2007a: Paoletti and Grulke. 2005; Bonn et al.. 2004; Kellomaki and
10 Wang. 1997; Tjoelker et al.. 1995; Reich and Lassoie. 1984). Since leaf-level stomatal
11 regulation is usually assessed in a steady state rather than as a dynamic response to
12 changing environmental conditions, steady state measurements cannot detect sluggish
13 stomatal response. Because of sluggish stomatal responses, water loss from plants may be
14 greater under dynamic environmental conditions over days and months.
15 In addition to the impacts on stomatal performance, O3-induced physiological changes,
16 such as reduced leaf area index and accelerated leaf senescence could alter water use
17 efficiency. It is well established from chamber and field studies that O3 exposure is
18 correlated with lower foliar retention (Karnosky et al.. 2003; Topaetal.. 2001; Pell et al..
19 1999; Grulke and Lee. 1997; Karnosky et al.. 1996; Miller et al.. 1972; Miller et al..
20 1963). However, Lee et al. (2009a) did not find changes in needle area of ponderosa pine
21 and reported that greater canopy conductance followed by water stress under elevated O3
22 may have been caused by stomatal dysfunction. At the Aspen FACE experiment, stand-
23 level water use, as indicated by sap flux per unit ground area, was not significantly
24 affected by elevated O3 despite a 22% decrease in leaf area index and 20% decrease in
25 basal area (Uddling et al.. 2008). The lack of negative effect of elevated O3 on stand
26 water use may be due to the substantially increased whole plant hydraulic conductance
27 per unit leaf area under elevated O3, as indicated by the sap flux per unit total leaf area
28 (kl) (Uddling et al.. 2009). The increased kl may be caused by the sluggish of stomatal
29 response. In pure aspen stands, the stomatal closure response to increasing vapor pressure
30 deficit was less sensitive and mid-day leaf water potential was lower under elevated O3,
31 suggesting O3 impaired stomatal control over transpiration (Uddling et al.. 2009). Other
32 potential factors contributing to the unchanged stand-level water use included the higher
33 proportion of sun leaves, and similar or even increased fine root biomass under elevated
34 O3 (Uddling et al.. 2008). Elevated O3 could also affect evapotranspiration by altering
35 tree crown interception of precipitation. Ozone has been shown to change branch
36 architectural parameters, and the effects were species-dependent at the Aspen FACE
37 experiment (Rheaet al.. 2010). The authors found that there was a significant correlation
38 between canopy architecture parameters and stem flow for birch but not aspen.
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1 It is difficult to scale up physiology measurements from leaves to ecosystems. Thus, the
2 current understanding of how stomatal response at leaf scale is integrated at the scale of
3 whole forest canopies, and therefore how it influences tree and forest stand water use is
4 limited. Field studies by McLaughlin et al. (2007a; 2007b) provided valuable insight into
5 the possible consequences of stomatal sluggishness for ecosystem water cycling.
6 McLaughlin et al. (2007a; 2007b) indicated that O3 increased water use in a mixed
7 deciduous forest in eastern Tennessee. McLaughlin et al. (2007a; 2007b) found that O3,
8 with daily maximum levels ranging from 69.2 to 82.9 ppb, reduced stem growth by 30-
9 50% in the high-O3 year 2002. The decrease in growth rate was caused in part by
10 amplification of diurnal cycles of water loss and recovery. Peak hourly O3 exposure
11 increased the rate of water loss through transpiration as indicated by the increased stem
12 sap flow. The authors suggested that a potential mechanism for the increased sap flow
13 could be altered stomatal regulation from O3 exposure, but this was inferred through sap
14 flow measurements and was not directly measured. The increased canopy water loss
15 resulted in higher water uptake by the trees as reflected in the reduced soil moisture in the
16 rooting zone. The change in tree water use led to further impacts on the hydrological
17 cycle at the landscape level. Increased water use under high O3 exposure was reported to
18 reduce late-season modeled streamflow in three forested watersheds in eastern Tennessee
19 (McLaughlin et al.. 2007b).
20 Felzer et al. (2009) used TEM-Hydro to assess the interactions of O3, climate, elevated
21 CO2 and N limitation on the hydrological cycle in the eastern U.S. They found that
22 elevated CO2 decreased evapotranspiration by 2-4% and increased runoff by 3-7%, as
23 compared to the effects of climate alone. When O3 damage and N limitation were
24 included, evapotranspiration was reduced by an additional 4-7% and runoff was increased
25 by an additional 6-11% (Felzer et al.. 2009). Based upon simulation with INTRAST and
26 LINKAGES, Hanson et al. (2005) found that increasing O3 concentration by 20 ppb
27 above the current ambient level yields a modest 3% reduction in water use. Those
28 ecological models were generally built on the assumption that O3 induces stomatal
29 closure and have not incorporated possible stomatal sluggishness due to O3 exposure.
30 Because of this assumption, results of those models normally found that O3 reduced
31 water use.
9.4.5.1 Summary
32 Although the evidence was from a limited number of field and modeling studies, findings
33 showed an association between O3 exposure and alteration of water use and cycling in
34 vegetation and at the ecosystem level. There is not a clear consensus on the nature of
35 leaf-level stomatal conductance response to O3 exposure. When measured under steady-
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1 state high light conditions, leaf-level stomatal conductance is often found to be reduced
2 when plants are exposed to O3. However, measurements of stomatal conductance under
3 dynamic light and VPD conditions indicate sluggish responses under elevated O3
4 exposure, which could potentially lead to increased water loss from vegetation. Field
5 studies conducted by McLaughlin et al. (2007a; 2007b) suggested that peak hourly O3
6 exposure increased the rate of water loss from several tree species, and led to a reduction
7 in the late-season modeled stream flow in three forested watersheds in eastern Tennessee.
8 Sluggish stomatal responses during O3 exposure was suggested as a possible mechanism
9 for increased water loss during peak O3 exposure. Currently, the O3-induced reduction in
10 stomatal aperture is the biological assumption for most process-based models. Because of
11 this assumption, results of those models normally found that O3 reduced water loss. For
12 example, Felzer (2009) found that O3 damage and N limitation together reduced
13 evapotranspiration and increased runoff.
14 Although the direction of the response differed among studies, the evidence is
15 sufficient to conclude that there is likely to be a causal relationship between O3
16 exposure and the alteration of ecosystem water cycling.
9.4.6 Below-Ground Processes
17 Above-ground and below-ground processes are tightly interconnected. Because roots and
18 soil organisms are not exposed directly to O3, below-ground processes are affected by O3
19 through alterations in the quality and quantity of C supply from photosynthates and
20 litterfall (Andersen. 2003). Ozone can decrease leaf C uptake by reducing photosynthesis
21 (Section 9.3). Ozone can also increase metabolic costs by stimulating the production of
22 chemical compounds for defense and repair processes, and by increasing the synthesis of
23 antioxidants to neutralize free radicals (see Section 9.3), both of which increase the
24 consumption of carbon for above-ground processes. Therefore, O3 could significantly
25 reduce the amount of C available for allocation to below-ground by decreasing C uptake
26 while increasing C consumption of above-ground processes (Andersen. 2003).
27 Since the 2006 O3 AQCD, there is additional evidence for O3 effects on below-ground
28 processes. Ozone has been found to alter root growth, soil food web structure,
29 decomposer activities, C turnover, water cycling and nutrient flow (Figure 9-8). Ozone
30 effects on root development and root biomass production and soil food web structure are
31 reviewed in sections 9.4.3.1 and 9.4.9.2, respectively. The focus in this section is on the
32 response of litter input, decomposer activities, soil respiration, soil C formation and
33 nutrient cycling.
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CO2, H2O CO2, H2O
Altered stomatal function
Litter production
and chemistry
CO, release
' Altered species competition
K "-1
Soil foodweb
•Bacteria
•Fungi
Micro & marco invertebrates
Soil physical &
chemical properties
Source: Modified from Andersen (2003)
Arrows denote C flux pathways that are affected by ozone. Dashed lines indicate where the impact of ozone is suspected but
unknown.
Figure 9-8 Conceptual diagram showing where ozone alters C, water and
nutrient flow in a tree-soil system, including transfer between biotic
and abiotic components below ground that influence soil physical
and chemical properties.
I
2
9.4.6.1 Litter Carbon Chemistry, Litter Nutrient and Their
Ecosystem Budgets
Consistent with previous findings, recent studies show that, although the responses are
often species-dependent, O3 tends to alter litter chemistry (U.S. EPA. 2006b).Alterations
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1 in chemical parameters, such as changes in C chemistry and nutrient concentrations, were
2 observed in both leaf and root litter (9-7).
3 At the Aspen FACE site, several studies investigated litter chemistry changes (Parsons et
4 al.. 2008; Johnson and Pregitzer. 2007; Chapman et al.. 2005; Liu et al.. 2005). In both
5 aspen and birch leaf litter, elevated O3 increased the concentrations of soluble sugars,
6 soluble phenolics and condensed tannins (Parsons et al.. 2008; Liu et al.. 2005).
7 Compared to other treatments, aspen litter under elevated O3 had the highest fiber
8 concentration, with the lowest concentration associated with the birch litter under the
9 same conditions (Parsons et al.. 2008). Chapman et al. (2005) measured chemical
10 changes in fine root litter and found that elevated O3 decreased lignin concentration. O3-
11 induced chemistry changes were also reported from other experimental sites. Results
12 from an OTC study in Finland suggested that elevated O3 increased the concentration of
13 acid-soluble lignin, but had no significant impact on other chemicals such as total sugars,
14 hemicelluloses, cellulose or total lignin in the litter of silver birch (Kasurinen et al..
15 2006). Results from the free air canopy O3 exposure experiment at Kranzberg Forest
16 showed that O3 increased starch concentrations but had no impact on cellulose and lignin
17 in beech and spruce leaf litter (Aneja et al.. 2007). The effect of O3 on three antioxidants
18 (ascorbate, glutathione and ot-tocopherol) in fine roots of beech was also assessed at
19 Kranzberg Forest. The results indicated that O3 had no significant effect on ot-tocopherol
20 and ascorbate concentrations, but decreased glutathione concentrations in fine roots
21 (Haberer et al., 2008). In addition to changing C chemistry, O3 also altered nutrient
22 concentrations in green leaves and litter (Table 9-6).
23 The combined effects of O3 on biomass productivity and chemistry changes may alter
24 C chemicals and nutrient contents at the canopy or ecosystem level. For example,
25 although O3 had different impacts on their concentrations, annual fluxes of C chemicals
26 (soluble sugar, soluble phenolics, condensed tannins, lipid and hemicelluloses), macro
27 nutrients (N, P, K and S) and micro nutrients (Mg, B, Cu and Zn) to soil were all reduced
28 due to lower litter biomass productivity at Aspen FACE (Liu et al., 2007a; Liu et al..
29 2005). At the Kranzberg Forest, N content of spruce canopy in a mixed culture and Ca2+
30 content of beech canopy in a monoculture increased due to elevated O3 increased leaf
31 concentrations of those nutrients although leaf production was not significantly altered by
32 O3 (Rodenkirchen et al.. 2009).
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Table 9-6 The effect of elevated ozone on leaf/litter nutrient concentrations
Study Site
Suonenjoki Research
Station, Finland
Aspen FACE
Aspen FACE
Kranzberg Forest, Germany
Kranzberg Forest, Germany
Salerno, Italy
Kuopio University Research
Garden, Finland
Species
Silver birch
Aspen and birch
Birch
Beech and spruce
Beech and spruce
Holm oak
Red Clover
Ozone
Concentration
Ambient: 10-60ppb
Elevated: 2xambient
Ambient: 50-60 ppb
Elevated: 1.5xambient
Ambient: 50-60 ppb
Elevated: 1.5xambient
Ambient: 9-41 ppb
Elevated: 2xambient
Ambient: 9-41 ppb
Elevated: 2xambient
Non-filtered OTC: 29 ppb
Filtered OTC:17ppb
Ambient: 25.7 ppb
Elevated: 1.5xambient
Response
Decreased the concentration of P, Mn, Zn
and B in leaf litter
Decreased the concentrations of P, S, Ca
and Zn, but had no impact on the
concentrations of N, K, Mg, Mn, B and Cu
in leaf litter.
Increase N concentration in birch litter
Increased N concentration in beach leaf,
but not in spruce needle
1) Had no significant effects on spruce
needle chemistry; 2) increased Ca
concentration in beech leaves in
monoculture, but had no impacts on other
nutrients
Ozone had no significant impacts on litter
C, N, lignin and cellulose concentrations
increased the total phenolic content of
leaves and had minor effects on the
concentrations of individual phenolic
compounds
Reference
Kasurinen et al.
(2006)
Liu et al. (2007a)
Parsons et al. (2008)
Kozovits et al. (2005)
Rodenkirchen etal.
(2009)
Baldantoni etal.
(2011)
Saviranta et
al.(2010)
1
2
3
4
5
6
9
10
11
12
13
14
15
16
17
9.4.6.2 Decomposer Metabolism and Litter Decomposition
The above- and below-ground physiological changes caused by O3 exposure cascade
through the ecosystem and affect soil food webs. In the 2006 O3 AQCD, there were very
few studies on the effect of O3 on the structure and function of soil food webs, except
two studies conducted by Larson et al. (2002) and Phillips et al. (2002). Since the last O3
AQCD, new studies have provided more information on how O3 affects the metabolism
of soil microbes and soil fauna.
Chung et al.(2.006) found that the activity of the cellulose-degrading enzyme 1,4-p-
glucosidase was reduced by 25% under elevated O3 at Aspen FACE. The decrease in
cellulose-degrading enzymatic activity was associated with the lower cellulose
availability under elevated O3 (Chung et al.. 2006). However, a later study at the same
site, which was conducted in the 10th year of the experiment, found that O3 had no
impact on cellulolytic activity in soil (Edwards and Zak. 2011). In a lysimeter study of
beech trees (Fagus sylvaticd) in Germany, soil enzyme activity was found to be
suppressed by O3 exposure (Esperschutz et al.. 2009; Pritsch et al.. 2009). Except for
xylosidase, enzyme activities involved in plant cell wall degradation (cellobiohydrolase,
beta-glucosidase and glucuronidase) were decreased in rhizosphere soil samples under
elevated O3 (2 x ambient level) (Pritsch etal.. 2009). Similarly, Chen et al. (2009) found
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1 O3 exposure, with a 3-month AOT40 of 21.4-44.1 ppm-h, decreased the microbial
2 metabolic capability in the rhizosphere and bulk soil of wheat, although the observed
3 reduction in bulk soil was not significant.
4 Ozone-induced change in soil organisms' activities could affect litter decomposition
5 rates. Results of recent studies indicated that O3 slightly reduced or have no impacts on
6 litter decomposition (Liu et al., 2009b; Parsons et al., 2008; Kasurinen et al., 2006)
7 (Baldantoni et al.. 2011). The responses varied among species, sites and exposure length.
8 Parsons et al. (2008) collected litter from aspen and birch seedlings at Aspen FACE site,
9 and conducted a 23-month field litter incubation starting in 1999. They found that
10 elevated O3 had different impacts on the decomposition of aspen and birch litter.
11 Elevated O3 was found to reduce aspen litter decomposition. However, O3 accelerated
12 birch litter decomposition under ambient CO2, but reduced it under elevated CO2
13 (Parsons et al.. 2008). Liu et al. (2009b) conducted another litter decomposition study at
14 Aspen FACE from 2003 to 2006, when stand leaf area index (LAI) reached its maximum.
15 During the 93 5-day field incubation, elevated O3 was shown to reduce litter mass loss in
16 the first year, but not in the second year. They suggested that higher initial tannin and
17 phenolic concentrations under elevated O3 reduced microbial activity in the first year
18 (Liu et al.. 2009b). In an OTC experiment, Kasurinen et al. (2006) collected silver birch
19 leaf litter from three consecutive growing seasons and conducted three separate litter-bag
20 incubation experiments. Litter decomposition was not affected by O3 exposure in the first
21 two incubations, but a slower decomposition rate was found in the third incubation. Their
22 principle component analysis indicated that the litter chemistry changes caused by O3
23 (decreased Mn, P, B and increased C:N) might be partially responsible for the decreased
24 mass loss of their third incubation. In another OTC experiment, Baldatoni et al. (2011)
25 found that O3 significantly reduced leaf litter decomposition of Quercus ilex L, although
26 litter C, N, lignin and cellulose concentrations were not altered by O3 exposure.
9.4.6.3 Soil respiration and carbon formation
27 Ozone could reduce the availability of photosynthates for export to roots, and increase
28 root mortality and turnover rates. Ozone has also been shown to reduce above-ground
29 litter productivity and alter litter chemistry, which would affect the quality and quantity
30 of the C supply to soil organisms (Section 9.4.6.1). The complex interactions among
31 those changes make it difficult to predict the response of soil C cycling under elevated
32 O3. The 2006 O3 AQCD concluded that O3 had no consistent impact on soil respiration
33 (U.S. EPA. 2006b). Ozone could increase or decrease soil respiration, depending on the
34 approach and timing of the measurements. Ozone may also alter soil C formation.
35 However, very few experiments directly measured changes in soil organic matter content
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1
2
3
4
5
6
under O3 fumigation (U.S. EPA. 2006b). Recent studies on soil respiration and soil
C content also found mixed responses. Most importantly, recent results from long-term
fumigation experiments, such as the Aspen FACE experiment, suggest that ecosystem
response to O3 exposure can change overtime. Observations made during the late
exposure years can be inconsistent with those during the early years, highlighting the
need for caution when assessing O3 effects based on short-term studies (Table 9-7).
Table 9-7 The temporal variation of ecosystem responses to ozone exposure
at Aspen FACE site
Endpoint
Litter decomposition
Fine root production
Soil respiration
Soil C formation
Period of
Measurement
1999-2001
2003-2006
1999
2002, 2005
1998-1999
2003-2007
1998-2001
2004-2008
Response
03 reduced aspen litter decomposition. However, 03 accelerated
birch litter decomposition under ambient C02, but reduced it under
elevated CO 2
03 reduced litter mass loss in the first year, but not in the second
year.
03 had no significant impact on fine root biomass
03 increased fine root biomass
Soil respiration under +C02+03 treatment was lower than that
under +C02 treatment
Soil respiration under +C02+03 treatment was 5-25% higher than
under elevated C02 treatment.
03 reduced the formation rates of total soil C by 51% and acid-
insoluble soil C by 48%
No significant effect of 03 on the new C formed under elevated
CO 2
Reference
Parsons et al. (2008)
Liu et al. (2009b)
Kina et al. (2001)
Pregitzeretal. (2008)
King etal. (2001)
Pregitzeretal. (2006) (2008)
Loya et al. (2003)
Talhelm etal. (2009)
9
10
11
12
13
14
15
16
17
18
Soil Respiration
Ozone has shown inconsistent impacts on soil respiration. A sun-lit controlled-
environment chamber study found that O3 had no significant effects on soil respiration,
fine root biomass or any of the soil organisms in a reconstructed ponderosa pine/soil-litter
system (Tingey et al.. 2006). In an adult European beech/Norway spruce forest at
Kranzberg Forest, the free air O3 fumigation (AOT40 of 10.2-117 ppm-h) increased soil
respiration under both beech and spruce during a humid year (Nikolova et al.. 2010) . The
increased soil respiration under beech has been accompanied by the increase in fine root
biomass and ectomycorrhizal fungi diversity and turnover (Grebenc and Kraigher. 2007).
The stimulating effect on soil respiration disappeared under spruce in a dry year, which
was associated with a decrease in fine root production in spruce under drought. This
finding suggested that drought was a more dominant stress than O3 for spruce (Nikolova
etal.. 2010). Andersen et al. (2010) labeled the canopies of European beech and Norway
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1 spruce with CO2 depleted in 13C at the same site. They did not observe any significant
2 changes in soil respiration for either species.
3 The nearly 10 year long studies at Aspen FACE indicated that the response of soil
4 respiration to O3 interacted with CO2 exposure and varied temporally (Table 9-7)
5 (Pregitzer et al. 2008; Pregitzeretal.. 2006; King etal.. 2001). Ozone treatment alone
6 generally had the lowest mean soil respiration rates, although those differences between
7 control and elevated O3 were usually not significant. However, soil respiration rates were
8 different with O3 alone and when acting in combination with elevated CO2. In the first
9 five years (1998-2002), soil respiration under +CO2+O3 treatment was similar to that
10 under control and lower than that under +CO2 treatment (Pregitzer et al., 2006; King et
11 al.. 2001). Since 2003, +CO2+O3 treatment started to show the greatest impact on soil
12 respiration. Compared to elevated CO2, soil respiration rate under +CO2+O3 treatment
13 was 15-25% higher from 2003-2004, and 5-10% higher from 2005-2007 (Pregitzer et al..
14 2008; Pregitzer et al., 2006). Soil respiration was highly correlated with the biomass of
15 roots with diameters of <2 mm and <1 mm, across plant community and atmospheric
16 treatments. The authors suggested that the increase in soil respiration rate may be due to
17 +CO2+O3 increased fine root (<1.0 mm) biomass production (Pregitzer et al.. 2008).
18 Changes in leaf chemistry and productivity due to O3 exposure have been shown to affect
19 herbivore growth and abundance (See Section 9.4.9.1). Canopy insects could affect soil
20 carbon and nutrient cycling through frass deposition, or altering chemistry and quantity
21 of litter input to the forest floor. A study at the Aspen FACE found that although elevated
22 O3 affected the chemistry of frass and greenfall, these changes had small impact on
23 microbial respiration and no effect on nitrogen leaching (Hillstrom et al.. 2010a).
24 However, respiratory carbon loss and nitrate immobilization were nearly double in
25 microcosms receiving herbivore inputs than those receiving no herbivore inputs
26 (Hillstrom et al.. 2010a).
Soil Carbon Formation
27 Ozone-induced reductions in plant growth can result in reduced C input to soil and
28 therefore soil C content (Andersen. 2003). The simulations of most ecosystem models
29 support this prediction (Ren et al., 2007a; Zhang et al.. 2007a; Felzer et al.. 2004).
30 However, very few studies have directly measured soil C dynamics under elevated O3.
31 After the first four years of fumigation (from 1998 to 2001) at the Aspen FACE site,
32 Loya et al. (2003) found that forest stands exposed to both elevated O3 and CO2
33 accumulated 51% less total soil C, and 48% less acid-insoluble soil C compared to stands
34 exposed only to elevated CO2. Soil organic carbon (SOC) was continuously monitored at
3 5 the Aspen FACE site, and the later data showed that the initial reduction in new
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1 C formation (soil C derived from plant litter since the start of the experiment) by O3
2 under elevated CO2 is only a temporary effect (Table 9-7) (Talhelm et al.. 2009). The
3 amount of new soil C in the elevated CO2 and the combined elevated CO2 and O3
4 treatments has converged since 2002. There was no significant effect of O3 on the new C
5 formed under elevated CO2 over the last four years of the study (2004-2008). Talhelm
6 et al. (2009) suggested the observed reduction in the early years of the experiment might
7 be driven by a suppression of C allocated to fine root biomass. During the early exposure
8 years, O3 had no significant impact on fine root production (King et al.. 2001). However,
9 the effect of O3 on fine root biomass was observed later in the experiment. Ozone
10 increased fine root production and the highest fine root biomass was observed under the
11 combined elevated CO2 and O3 treatment in the late exposure years (Table 9-7)
12 (Pregitzer et al.. 2006). This increase in fine root production was due to changes in
13 community composition, such as better survival of O3-tolerant aspen genotype, birch and
14 maple, rather than changes in C allocation at the individual tree level (Pregitzer et al..
15 2008; Zak et al.. 2007).
9.4.6.4 Nutrient cycling
16 Ozone can affect nutrient cycling by changing nutrient release from litter, nutrient uptake
17 by plants, and soil microbial activity. Nitrogen is the limiting nutrient for most temperate
18 ecosystems, and several studies examined N dynamics under elevated O3. Nutrient
19 mineralization from decomposing organic matter is important for sustaining ecosystem
20 production. Holmes et al. (2006) found that elevated O3 decreased gross N mineralization
21 at the Aspen FACE site, indicating that O3 may reduce N availability. Other N cycling
22 processes, such as NH4+ immobilization, gross nitrification, microbial biomass N and soil
23 organic N, were not affected by elevated O3 (Holmes et al.. 2006). Similarly, Kanerva
24 et al. (2006) found total N, NO3-, microbial biomass N, potential nitrification and
25 denitrification in their meadow mesocosms were not affected by elevated O3 (40-50 ppb).
26 Ozone was found to decreased soil mineral N content at SoyFACE, which was likely
27 caused by a reduction in plant material input and increased denitrification (Pujol Pereira
28 et al.. 2011). Ozone also showed small impact on other micro and macro nutrients. Liu
29 et al. (2007a) assessed N, P, K, S, Ca, Mg, Mn, B, Zn and Cu release dynamics at Aspen
30 FACE, and they found that O3 had no effects on most nutrients, except to decrease N and
31 Ca release from litter. These studies reviewed above suggested that soil N cycling
32 processes were not affected or slightly reduced by O3 exposure. However, in a lysimeter
33 study with young beech trees Stoelken et al. (2010) found that elevated O3 stimulated N
34 release from litter which was largely attributed to an enhanced mobilization of inert
35 nitrogen fraction.
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1 Using the Simple Nitrogen Cycle model (SINIC), Hong et al. (2006) evaluated the
2 impacts of O3 exposure on soil N dynamics and streamflow nitrate flux. The detrimental
3 effect of O3 on plant growth was found to reduce plant uptake of N and therefore increase
4 nitrate leaching. Their model simulation indicated that ambient O3 exposure increased the
5 mean annual stream flow nitrate export by 12% (0.042 g N/m2/year) at the Hubbard
6 Brook Experimental Watershed from 1964-1994 (Hong et al.. 2006).
9.4.6.5 Dissolved Organic Carbon and Biogenic Trace Gases
Emission
7 The O3-induced changes in plant growth, C and N fluxes to soil and microbial
8 metabolism can alter other biogeochemical cycling processes, such as soil dissolved
9 organic carbon (DOC) turnover and trace gases emission.
10 Jones et al. (2009) collected fen cores from two peatlands in North Wales, UK and
11 exposed them to one of four levels of O3 (AOT40 of 0, 3.69, 5.87 and 13.80 ppm-h for
12 41 days). They found the concentration of porewater DOC in fen cores was significantly
13 decreased by increased O3 exposure. A reduction of the low molecular weight fraction of
14 DOC was concurrent with the observed decrease in DOC concentration. Their results
15 suggested that O3 damage to overlying vegetation may decrease utilizable C flux to soil.
16 Microbes, therefore, have to use labile C in the soil to maintain their metabolism, which,
17 the authors hypothesized, leads to a decreased DOC concentration with a shift of the
18 DOC composition to more aromatic, higher molecular weight organic compounds.
19 Several studies since the 2006 O3 AQCD have examined the impacts of O3 on nitrous
20 oxide (N2O) and methane (CH4) emission. Kanerva et al. (2007) measured the fluxes of
21 N2O and CH4 in meadow mesocosms, which were exposed to elevated CO2 and O3 in
22 OTCs in south-western Finland. They found that the daily N2O fluxes were decreased in
23 the NF+O3 (non-filtered air + elevated O3, 40-50 ppb) after three seasons of exposure.
24 Elevated O3 alone or combined with CO2 did not have any significant effect on the daily
25 fluxes of CH4 (Kanerva et al.. 2007). In another study conducted in central Finland, the
26 4 year open air O3 fumigation (AOT40 of 20.8-35.5 ppm-h for growing season) slightly
27 increased potential CH4 oxidation by 15% in the peatland microcosms, but did not affect
28 the rate of potential CH4 production or net CH4 emissions, which is the net result of the
29 potential CH4 production and oxidation (Morsky et al.. 2008). However, several studies
30 found that O3 could significantly reduce CH4 emission. Toet et al. (2011) exposed
31 peatland mesocosms to O3 in OTCs for two years, and found that CH4 emissions were
32 significantly reduced by about 25% during midsummer periods of both years. In an OTC
33 study of rice paddy, Zheng et al. (2011) found that the daily mean CH4 emissions were
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1 significantly lower under elevated O3 treatments than those in charcoal-filtered air and
2 nonfiltered air treatments. They found that the seasonal mean CH4 emissions were
3 negatively related with AOT40, but positively related to the relative rice yield,
4 aboveground biomass and underground biomass.
9.4.6.6 Summary
5 Since the 2006 O3 AQCD, more evidence has shown that although the responses are
6 often site specific, O3 altered the quality and quantity of litter input to soil, microbial
7 community composition, and C and nutrient cycling. Biogeochemical cycling of below-
8 ground processes is driven by C input from plants. Studies at the leaf and plant level have
9 provided biologically plausible mechanisms, such as reduced photosynthetic rates,
10 increased metabolic cost, and reduced root C allocation for the association of O3
11 exposure and the alteration of below-ground processes.
12 Results from Aspen FACE and other experimental studies consistently found that O3
13 reduced litter production and altered C chemistry, such as soluble sugars, soluble
14 phenolics, condensed tannins, lignin, and macro/micro nutrient concentration in litter
15 (Parsons et al., 2008; Kasurinen et al., 2006; Liu et al., 2005). The changes in substrate
16 quality and quantity could alter microbial metabolism under elevated O3, and therefore
17 soil C and nutrient cycling. Several studies indicated that O3 suppressed soil enzyme
18 activities (Pritsch et al.. 2009; Chung et al.. 2006). However, the impact of O3 on litter
19 decomposition was inconsistent and varied among species, sites and exposure length.
20 Similarly, O3 had inconsistent impacts on dynamics of micro and macro nutrients.
21 Studies from the Aspen FACE experiment suggested that the response of below-ground
22 C cycle to O3 exposure, such as litter decomposition, soil respiration and soil C content,
23 changed over time. For example, in the early part of the experiment (1998-2003), O3 had
24 no impact on soil respiration but reduced the formation rates of total soil C under
25 elevated CO2. However, after 10-11 yr of exposure, O3 was found to increase soil
26 respiration but have no significant impact on soil C formation under elevated CO2.
27 The evidence is sufficient to infer that there is a causal relationship between O3
28 exposure and the alteration of below-ground biogeochemical cycles.
9.4.7 Community composition
29 The effects of O3 on species competition (AX9.3.3.4) and community composition
30 (AX9.6.4) were summarized in the 2006 O3 AQCD. Plant species differ in their
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1 sensitivity to O3. Fast growing plants with high stomatal conductance and high specific
2 leaf area (SLA) were more likely to be sensitive to O3 exposure. Further, different
3 genotypes of a given species also vary in their sensitivity. This differential sensitivity
4 could change the competitive interactions that lead to loss in O3 sensitive species or
5 genotypes. In addition, O3 exposure has been found to alter reproductive processes in
6 plants (See Section 9.4.3.3). Changes in reproductive success could lead to changes in
7 species composition. However, since ecosystem-level responses result from the
8 interaction of organisms with one another and with their physical environment, it takes
9 longer for a change to develop to a level of prominence at which it can be identified and
10 measured. A shift in community composition in forest and grassland ecosystems noted in
11 the 2006 O3 AQCD has continued to be observed from experimental and gradient studies.
12 Additionally, research since the last review has shown that O3 can alter community
13 composition and diversity of soil microbial communities.
9.4.7.1 Forest
14 In the San Bernardino Mountains in southern California, O3 pollution caused a
15 significant decline in ponderosa pine (Pinus ponderosa ) and Jeffrey pine (Pinus jeffreyi}
16 (U.S. EPA. 2006b). Pine trees in the young mature age class group exhibited higher
17 mortality rates compared with mature trees at a site with severe O3 visible foliar injury.
18 The vulnerability of young mature pines was most likely caused by the fact that trees in
19 this age class were emerging into the canopy, where higher O3 concentrations were
20 encountered (McBride and Laven. 1999). Because of the loss of O3-sensitive pines,
21 mixed forests of ponderosa pine, Jeffery Pine and white fir (Abies concolor) shifted to
22 predominantly white fir (Miller, 1973). Ozone may have indirectly caused the decline in
23 understory diversity in coniferous forests in the San Bernardino Mountains through an
24 increase in pine litterfall. This increase in litterfall from O3 exposure results in an
25 understory layer that may prohibit the establishment of native herbs, but not exotic annual
26 Galium aparine (Allen et al.. 2007).
27 Ozone damage to conifer forests has also been observed in several other regions. In the
28 Valley of Mexico, a widespread mortality of sacred fir (Abies religiosd) was observed in
29 the heavily polluted area of the Desierto de los Leones National Park in the early 1980s
30 (de Lourdes de Bauer and Hernandez-Tejeda, 2007; Fenn et al.. 2002). Ozone damage
31 was widely believed to be an important causal factor in the dramatic decline of sacred fir.
32 In alpine regions of southern France and the Carpathians Mountains, O3 was also
33 considered as the major cause of the observed decline in cembran pine (Pinus
34 cembra)(Wieser et al., 2006). However, many environmental factors such as light,
35 temperature, nutrient and soil moisture, and climate extremes such as unusual dry and
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1 wet periods could interact with O3 and alter the response of forest to O3 exposure. For
2 those pollution gradient studies, several confounding factors, such as drought, insect
3 outbreak and forest management, may also contribute to or even be the dominant factors
4 causing the mortality of trees (de Lourdes de Bauer and Hernandez-Tejeda. 2007; Wieser
5 et al.. 2006).
6 New evidence from long-term free O3 fumigation experiments provided additional
7 support for the potential impacts of O3 on species competition and community
8 composition changes in forest ecosystems. At the Aspen FACE site, community
9 composition at both the genetic and species levels was altered after seven years of
10 fumigation with O3 (Kubiske et al., 2007). In the pure aspen community, O3 fumigation
11 reduced growth and increased mortality of sensitive clone 259, while the O3 tolerant
12 clone 8L emerged as the dominant clone. Growth of clone 8L was even greater under
13 elevated O3 compared to controls, probably due to O3 alleviated competitive pressure on
14 clone 8L by reducing growth of other clones. In the mixed aspen-birch and aspen-maple
15 communities, O3 reduced the competitive capacity of aspen compared to birch and maple
16 (Kubiske et al.. 2007). In a phytotron study, O3 fumigation reduced growth of beech but
17 not spruce in mixed culture, suggesting a higher susceptibility of beech to O3 under
18 interspecific competition (Kozovits et al.. 2005).
9.4.7.2 Grassland and Agricultural Land
19 The response of managed pasture, often cultivated as a mixture of grasses and clover, to
20 O3 pollution has been studied for many years. The tendency for O3-exposure to shift the
21 biomass of grass-legume mixtures in favor of grass species, reported in the previous O3
22 AQCD has been generally confirmed by recent studies. In a mesocosm study, Trifolium
23 repens and Loliumperenne mixtures were exposed to an episodic rural O3 regime within
24 solardomes for 12 weeks. T. repens showed significant changes in biomass but notZ.
25 perenne, and the proportion of T. repens decreased in O3-exposed mixtures compared to
26 the control (Haves et al., 2009). The changes in community composition of grass-legume-
27 forb mixtures were also observed at the Le Mouret FACE experiment, Switzerland.
28 During the 5-year O3 fumigation (AOT40 of 13.3-59.5 ppm-h), the dominance of
29 legumes in fumigated plots declined more quickly than those in the control plots (Yolk et
30 al., 2006). However, Stampfli and Fuhrer (2010) re-analyzed the species and soil data and
31 suggested that Volk et al. (2006) overestimated the O3 effect. Stampfli and Fuhrer (2010)
32 found that the difference in the species dynamics between control and O3 treatment was
33 more caused by heterogeneous initial conditions than O3 exposure. Several studies also
34 suggested the mature/species-rich ecosystems were more resilient to O3 exposure. At
35 another FACE experiment, located at Alp Flix, Switzerland, O3 fumigation (AOT40 of
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1 15.2-64.9 ppm-h) showed no significant impact on community composition of this
2 species-rich pasture (Bassin et al.. 2007b). Although most studies demonstrated an
3 increase in grass:forb ration with O3 exposure (Haves et al., 2009; U.S. EPA. 2006b). a
4 study on a simulated upland grassland community O3 reduce grass:forb ratio (Felicity et
5 al., 2010). which may be due to grass species in this community, such as Anthoxanthum
6 odoratum, was more sensitive to O3 than other most studied grass species such as L.
7 perenne (Haves et al., 2009). Pfleeger et al. (2010) collected seed bank soil from an
8 agricultural field and examined how the plant community responded over several
9 generations to elevated O3 exposures. Sixty plant species from 22 families emerged in the
10 chambers over their four year study. Overall, they found that O3 appeared to have small
11 impacts on seed germination and only a minor effect on species richness of pioneer plant
12 communities.
13 Several review papers have discussed the physiological and ecological characteristics of
14 O3-sensitive herbaceous plants. Hayes et al. (2007) assessed species traits associated with
15 O3 sensitivity by the changes in biomass caused by O3 exposure. Plants of the therophyte
16 (e.g., annual) life form were particularly sensitive to O3. Species with higher mature leaf
17 N concentration tended to be more sensitive than those with lower leaf N concentration.
18 Plants growing under high oxidative stress environments, such as high light or high
19 saline, were more sensitive to O3. Using the same dataset from Hayes et al. (2007). Mills
20 et al. (2007a) identified the O3 sensitive communities. They found that the largest number
21 of these O3 sensitive communities were associated with grassland ecosystems. Among
22 grassland ecosystems, alpine grassland, sub-alpine grassland, woodland fringe, and dry
23 grassland were identified as the most sensitive communities.
9.4.7.3 Microbes
24 Several methods have been used to study microbial composition changes associated with
25 elevated O3. Phospholipid fatty acid (PLFA) analysis is widely used to determine
26 whether O3 elicits an overall effect on microbial community composition. However,
27 since PLFA markers cover a broad range of different fungi, resolution of this method
28 may be not fine enough to detect small changes in the composition of fungal
29 communities. Methods, such as microscopic analyses and polymerase chain reaction-
30 denaturing gradient gel electrophoresis (PCR-DGGE), have better resolution to
31 specifically analyze the fungal community composition. The resolution differences
32 among those methods needs to be considered when assessing the O3 impact on microbial
33 community composition.
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1 Kanerva et al. (2008) found that elevated O3 (40-50 ppb) decreased total, bacterial,
2 actinobacterial and fungal PLFA biomass values as well as fungal:bacterial PLFA
3 biomass ratio in their meadow mesocosms in south-western Finland. The relative
4 proportions of individual PLFAs between the control and elevated O3 treatments were
5 significantly different, suggesting that O3 modified the structure of the microbial
6 community. Morsky et al. (2008) exposed boreal peatland microcosms to elevated O3,
7 with growing season AOT40 of 20.8-35.3 ppm-h, in an open-air O3 exposure field in
8 Central Finland. They also found that microbial composition was altered after three
9 growing seasons with O3 fumigation, as measured by PLFA. Ozone tended to increase
10 the presence of Gram-positive bacteria and the biomass of fungi in the peatland
11 microcosms. Ozone also resulted in higher microbial biomass, which co-occurred with
12 the increases in concentrations of organic acids and leaf density of sedges (Morsky et al..
13 2008). In a lysimeter experiment in Germany, O3 was found to alter the PLFA profiles in
14 the upper 0-20 cm rhizosphere soil of European beech. Elevated O3 reduced bacterial
15 abundance but had no detectable effect on fungal abundance (Pritsch et al.. 2009). Using
16 microscopic analyses, Kasurinen et al. (2005) found that elevated O3, with 5 or 6 months
17 of AOT40 of 20.6-30.9 ppm-h, decreased the proportions of black and liver-brown
18 mycorrhizas and increased that of light brown/orange mycorrhizas. In an herbaceous
19 plant study, SSCP (single-strand conformation polymorphism) profiles indicated that O3
20 stress (about 75 ppb) had a very small effect on the structural diversity of the bacterial
21 community in rhizospheres (Dohrmann and Tebbe. 2005). At the Aspen FACE site, O3
22 had no significant effect on fungal relative abundance, as indicated by PLFA profile.
23 However, elevated O3 altered fungal community composition, according to the
24 identification of 39 fungal taxonomic units from soil using polymerase chain reaction-
25 denaturing gradient gel electrophoresis (PCR-DGGE) (Chung et al.. 2006). In another
26 study at Aspen FACE, phylogenetic analysis suggested that O3 exposure altered
27 agaricomycete community. The ectomycorrhizal communities developing under elevated
28 O3 had higher proportions of Cortinarius and Inocybe species, and lower proportions of
29 Laccaria and Tomentella (Edwards and Zak. 2011). Ozone was found to change
30 microbial community composition in an agricultural system. Chen et al. (201 Ob) found
31 elevated O3 (100-150 ppb) had significant effects on soil microbial composition
32 expressed as PLFA percentage in a rice paddy in China.
9.4.7.4 Summary
33 In the 2006 O3 AQCD, the impact of O3 exposure on species competition and community
34 composition was assessed. Ozone was found to cause a significant decline in ponderosa
3 5 and Jeffrey pine in the San Bernardino Mountains in southern California. Ozone exposure
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1 also tended to shift the grass-legume mixtures in favor of grass species (U.S. EPA.
2 20061)). Since the 2006 O3 AQCD, more evidence has shown that O3 exposure changed
3 the competitive interactions and could lead to loss of O3 sensitive species or genotypes.
4 Studies at plant level found that the severity of O3 damage on growth, reproduction and
5 foliar injury varied among species, which provided the biological plausibility for the
6 alteration of community composition. Additionally, research since the last review has
7 shown that O3 can alter community composition and diversity of soil microbial
8 communities.
9 The decline of conifer forests under O3 exposure was continually observed in several
10 regions. Ozone damage was believed to be an important causal factor in the dramatic
11 decline of sacred fir in the valley of Mexico (de Lourdes de Bauer and Hernandez-
12 Tejeda. 2007). as well as cembran pine in southern France and Carpathian Mountains
13 ("Wieser et al.. 2006). Results from the Aspen FACE site indicated that O3 could alter
14 community composition of broadleaf forests as well. At the Aspen FACE site, O3
15 reduced growth and increased mortality of a sensitive aspen clone, while the O3 tolerant
16 clone emerged as the dominant clone in the pure aspen community. In the mixed aspen-
17 birch and aspen-maple communities, O3 reduced the competitive capacity of aspen
18 compared to birch and maple (Kubiske et al.. 2007).
19 The tendency for O3-exposure to shift the biomass of grass-legume mixtures in favor of
20 grass species, was reported in the 2006 O3 AQCD and has been generally confirmed by
21 recent studies. However, in a high elevation mature/species-rich grass-legume pasture, O3
22 fumigation showed no significant impact on community composition (Bassin et al..
23 2007b).
24 Ozone exposure not only altered community composition of plant species, but also
25 microorganisms. The shift in community composition of bacteria and fungi has been
26 observed in both natural and agricultural ecosystems, although no general patterns could
27 be identified OCanerva et al.. 2008; Morsky et al.. 2008; Kasurinen et al.. 2005).
28 The evidence is sufficient to conclude that there is likely a causal relationship
29 between O3 exposure and the alteration of community composition.
9.4.8 Factors that Modify Functional and Growth Response
30 Many biotic and abiotic factors, including insects, pathogens, root microbes and fungi,
31 temperature, water and nutrient availability, and other air pollutants, as well as elevated
32 CO2, influence or alter plant response to O3. These modifying factors were
33 comprehensively reviewed in AX9.3 of the 2006 O3 AQCD and thus, this section serves
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1 mainly as a brief summary of the previous findings. A limited number of new studies
2 published since the 2006 O3 AQCD add to our understanding of the role of these
3 interactions in modifying O3-induced plant responses. Many of these modifying factors
4 and interactions are integrated into discussions elsewhere in this chapter and the reader is
5 directed to those sections.
9.4.8.1 Genetics
6 It is well known that species vary greatly in their responsiveness to O3. Even within a
7 given species, individual genotypes or populations can also vary significantly with
8 respect to O3 sensitivity (U.S. EPA. 2006b). Therefore, caution should be taken when
9 considering a species' degree of sensitivity to O3. Plant response to O3 is determined by
10 genes that are directly related to oxidant stress and to an unknown number of genes that
11 are not specifically related to oxidants, but instead control leaf and cell wall thickness,
12 stomatal conductance, and the internal architecture of the air spaces. It is rarely the case
13 that single genes are responsible for O3 tolerance. Studies using molecular biological
14 tools and transgenic plants have positively verified the role of various genes and gene
15 products in O3 tolerance and are continuing to increase the understanding of O3 toxicity
16 and differences in O3 sensitivity. See Section 9.3.3.2 of this document for a discussion of
17 recent studies related to gene expression changes in response to O3.
9.4.8.2 Environmental Biological Factors
18 As stated in the 2006 O3 AQCD, the biological factors within the plant's environment
19 that may influence its response to O3 encompass insects and other animal pests, diseases,
20 weeds, and other competing plant species. Ozone may influence the severity of a disease
21 or infestation by a pest or weed, either by direct effects on the causal species, or
22 indirectly by affecting the host, or both. In addition, the interaction between O3, a plant,
23 and a pest, pathogen, or weed may influence the response of the target host species to O3
24 (U.S. EPA. 2006b). Several recent studies on the effects of O3 on insects via their
25 interactions with plants are discussed in Section 9.4.9.1. In addition, O3 has also been
26 shown to alter soil fauna communities (Section 9.4.9.2).
27 In contrast to detrimental biological interactions, there are mutually beneficial
28 relationships or symbioses involving higher plants and bacteria or fungi. These include
29 (1) the nitrogen-fixing species Rhizobium and Frankia that nodulate the roots of legumes
30 and alder and (2) the mycorrhizae that infect the roots of many crop and tree species, all
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1 of which may be affected by exposure of the host plants to O3. Some discussion of
2 mycorrhizae can be found in Section 9.4.6.
3 In addition to the interactions involving animal pests, O3 also has indirect effects on
4 higher herbivorous animals, e.g., livestock, due to O3-induced changes in feed quality.
5 Recent studies on the effects of O3 on nutritive quality of plants are discussed in Sections
6 9.4.4.2.
7 Intra- and interspecific competition are also important factors in determining vegetation
8 response to O3. Plant competition involves the ability of individual plants to acquire the
9 environmental resources needed for growth and development: light, water, nutrients, and
10 space. Intraspecific competition involves individuals of the same species, typically in
11 monoculture crop situations, while interspecific competition refers to the interference
12 exerted by individuals of different species on each other when they are in a mixed
13 culture. This topic was previously reviewed in AX9.3.3.4 of the 2006 O3 AQCD. Recent
14 studies on competition and its implications for community composition are discussed in
15 Section 9.4.7.
9.4.8.3 Physical Factors
16 Physical or abiotic factors play a large role in modifying plant response to O3, and have
17 been extensively discussed in previous O3 AQCDs. This section summarizes those
18 findings as well as recent studies published since the last review.
19 Although some studies have indicated that O3 impact significantly increases with
20 increased ambient temperature (Ball et al.. 2000; Mills et al.. 2000). other studies have
21 indicated that temperature has little effect (Balls et al.. 1996; Fredericksen et al.. 1996). A
22 recent study by Riikonen et al. (2009) at the Ruohoniemi open air exposure field in
23 Kuopio, Finland found that the effects of temperature and O3 on total leaf area and
24 photosynthesis of Betulapendula were counteractive. Elevated O3 reduced the saplings'
25 ability to utilize the warmer growth environment by increasing the stomatal limitation for
26 photosynthesis and by reducing the redox state of ascorbate in the apoplast in the
27 combination treatment as compared to temperature alone (Riikonen et al.. 2009).
28 Temperature affects the rates of all physiological processes based on enzyme catalysis
29 and diffusion; each process and overall growth (the integral of all processes) has a
30 distinct optimal temperature range. It is important to note that a plant's response to
31 changes in temperature will depend on whether it is growing near its optimum
32 temperature for growth or near its maximum temperature (Rowland-Bamford. 2000).
33 However, temperature is very likely an important variable affecting plant O3 response in
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1 the presence of the elevated CO2 levels contributing to global climate change. In contrast,
2 some evidence suggests that O3 exposure sensitizes plants to low temperature stress
3 (Colls and Unsworth. 1992) and, also, that O3 decreases below-ground carbohydrate
4 reserves, which may lead to responses in perennial species ranging from rapid demise to
5 impaired growth in subsequent seasons (i.e., carry-over effects) (Andersen et al., 1997).
6 Light, a component of the plant's physical environment, is an essential "resource" of
7 energy content that drives photosynthesis and C assimilation. It has been suggested that
8 increased light intensity may increase the O3 sensitivity of light-tolerant species while
9 decreasing that of shade-tolerant species, but this appears to be an oversimplification with
10 many exceptions. Several studies suggest that the interaction between O3 sensitivity and
11 light environment is complicated by the developmental stage as well as the light
12 environment of individual leaves in the canopy (Kitao et al., 2009; Topaet al., 2001;
13 Chappelkaand Samuelson. 1998).
14 Although the relative humidity of the ambient air has generally been found to increase the
15 effects of O3 by increasing stomatal conductance (thereby increasing O3 flux into the
16 leaves), abundant evidence also indicates that the ready availability of soil moisture
17 results in greater O3 sensitivity (Mills. 2002). The partial "protection" against the effects
18 of O3 afforded by drought has been observed in field experiments (Low et al., 2006) and
19 modeled in computer simulations (Broadmeadow and Jackson. 2000). Conversely,
20 drought may exacerbate the effects of O3 on plants (Pollastrini et al., 2010; Grulke et al.,
21 2003b). There is also some evidence that O3 can predispose plants to drought stress
22 (Maier-Maercker. 1998). Hence, the nature of the response is largely species-specific and
23 will depend to some extent upon the sequence in which the stressors occur.
9.4.8.4 Interactions with other Pollutants
Ozone-Nitrogen Interactions
24 Elevated O3 exposure and N deposition often co-occur. However, the interactions of O3
25 exposure and N deposition on vegetation are complex and less well understood compared
26 to their independent effects. Consistent with the conclusion of the 2006 O3 AQCD, the
27 limited number of studies published since the last review indicated that the interactive
28 effects of N and O3 varied among species and ecosystems (Table 9-8). To better
29 understand these interactions in ecosystems across the U.S., more information is needed
30 considering combined O3 exposure and N deposition related effects.
31 Nitrogen deposition could stimulate relative growth rate (RGR), and lead to increased
32 stomatal conductance. Therefore, plants might become more susceptible to O3 exposure.
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1 Alternatively, N deposition may increase the availability of photosynthates for use in
2 detoxification and plants could become more tolerant to O3 (Bassin et al.. 2007a). Only a
3 few recent studies have investigated the interactive effects of O3 and N in the U.S. Grulke
4 et al. (2005) measured stomatal conductance of California black oak (Quercus kelloggii)
5 at a long-term N-enrichment site located in the San Bernardino Mountains, which is
6 accompanied by high O3 exposure (80 ppb, 24-h avg. over a six month growing season).
7 The authors found that N amendment led to poor stomatal control in full sun in
8 midsummer of the average precipitation years, but enhanced stomatal control in shade
9 leaves of California black oak. In an OTC study, Handley and Grulke (2008) found that
10 O3 lowered photosynthetic ability and water-use efficiency, and increased leaf chlorosis
11 and necrosis of California black oak. Nitrogen fertilization tended to reduce plant
12 sensitivity to O3 exposure; however, the interaction was not statistically significant.
13 Studies conducted outside the U.S. are also summarized in Table 9-8. Generally, the
14 responses were species specific. The O3-induced reduction in photosynthetic rate and
15 biomass loss were greater in the relatively high N treatment for watermelon (Citrillus
16 tenants) (Calatayud et al., 2006) and Japanese beech (Fagus crenata) seedlings
17 (Yamaguchi et al.. 2007). However, there was no significant interactive effect of O3 and
18 N on biomass production for Quercus serrata seedlings (Watanabe et al.. 2007). young
19 Norway spruce (Picea abies) trees (Thomas et al.. 2005). and young European beech
20 (Fagus sylvatica) trees (Thomas et al.. 2006).
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Table 9-8 Response of plants to the interactive effects of elevated ozone
exposure and N enrichment
Site
San
Bernardino
Mountains,
U.S.
San
Bernardino
Mountains,
U.S.
Switzerland
Switzerland
Switzerland
Switzerland
Switzerland
Switzerland
Spain
Spain
Japan
Japan
Species
California black
oak (Quercus
kelloggii)
California black
oak (Quercus
kelloggii)
Spruce trees
(P/cea abies)
Beech trees
(Fagus sylvatica)
Alpine pasture
Alpine pasture
Alpine pasture
Alpine pasture
Watermelon
(Citrillus tenants)
Trifolium striatum
Japanese beech
seedlings (Fagus
crenafa)
Quercus serrata
seedlings
Ozone exposure
SOppb
0,75, and 1 SOppb
Filtered (19.4-28.1 ppb);
ambient (37.6-47.4 ppb)
Filtered (19.4-28.1 ppb);
ambient (37.6-47.4 ppb)
Ambient (AOT40 of 11.1-
12.6ppm-h); 1.2 ambient
(AOT40 of 15.2 -29.5 ppm-
h) and 1 .6 ambient (28.4-
64.9 ppm-h)
Ambient (AOT40 of 11.1-
12.6 ppm-h); 1.2 ambient
(AOT40 of 15.2-29.5 ppm-
h) and 1 .6 ambient (28.4-
64.9 ppm-h)
Ambient (AOT40 of 11.1-
12.6 ppm-h); 1.2 ambient
(AOT40 of 15.2-29.5 ppm-
h) and 1 .6 ambient (28.4-
64.9 ppm-h)
Ambient (AOT40 of 11.1-
12.6 ppm-h); 1.2 ambient
(AOT40 of 15.2-29.5 ppm-
h) and 1 .6 ambient (28.4-
64.9 ppm-h)
03free (AOT40 of 0 ppm-
h), ambient (AOT40 of 5.1-
6.3 ppm-h) and elevated
03(AOT40 of 32.5-35.6
ppm-h)
Filtered (24-h avg. of 8-22
ppb); ambient (29-34 ppb),
elevated 0 3 (35-56 ppb)
Filtered (24-h avg. of 10.3-
13.2 ppb); ambient (42.0-
43.3 ppb), 1.5 ambient
(62.6-63.9 ppb) and 2.0
ambient (82.7-84.7 ppb)
Filtered (24-h avg. of 10.3-
13.2 ppb); ambient (42.0-
43.3 ppb), 1.5 ambient
(62.6-63.9 ppb) and 2.0
ambient (82.7-84.7 ppb)
N addition
0, and 50 kg N/
ha/yr
0, and 50 kg N/
ha/yr
0, 20, 40 and 80
kg N/ ha/yr
0, 20, 40 and
80 kg N/ ha/yr
0,5, 10' 25, 50
kg N/ ha/yr
0,5,10,25,50
kg N/ha/yr
0,5, 10' 25, 50
kg N/ ha/yr
0,5, 10' 25, 50
kg N/ ha/yr
140, 280, and
436 kg N/ ha/yr
10, 30, and 60
kg N/ ha/yr
0, 20 and 50 kg
N/ ha/yr
0, 20 and 50 kg
N/ ha/yr
Responses
N-amended trees had lower late
summer C gain and greater foliar
chlorosis in the drought year, and poor
stomatal control and lower leaf water
use efficiency and in midsummer of the
average precipitation year.
N fertilization tended to reduce plant
sensitivity to 03 exposure; however
the interaction was not statistically
significant.
Higher N levels alleviated the negative
impact of 03 on root starch concentrations
N addition amplified the negative
effects of 0 3 on leaf area and shoot
elongation.
The positive effects of N addition on
canopy greenness were counteracted
by accelerated leaf senescence in the
highest 03 treatment.
Only a small number of species
showed significant 03 and N
interactive effects on leaf chlorophyll
concentration, leaf weight and change
in 180, and the patterns were not
consistent.
The positive effects of N addition on
canopy greenness were counteracted
by accelerated leaf senescence in the
highest 03 treatment.
Highest N addition resulted in carbon
loss, but there was no interaction
between 03 and N treatments.
High N concentration enhanced the
detrimental effects of 03 on
Chlorophyll a fluorescence
parameters, lipid peroxidation, and the
total yield.
03 reduced total aerial biomass. N
fertilization counterbalanced 03-
induced effects only when plants were
exposed to moderate 03 levels
(ambient) but not under elevated 03
concentrations.
The 03-induced reduction in net
photosynthesis and whole-plant dry
mass were greater in the relatively
high N treatment than that in the low N
treatment.
No significant interactive effects of 03
and N load on the growth and net
photosynthetic rate were detected.
References
Grulke et al. (2005)
Handley andGruIke
(2008)
Thomas et al. (2005)
Thomas et al. (2006)
Bassin et al. (2007b)
Bassin et al. (2009)
Bassin et al. (2007b)
Volketal. (2011)
Calatayud et al. (2006)
Sanz et al. (2007)
Yamaguchi etal.
(2007)
Watanabe etal. (2007)
1
2
Ozone-Carbon Dioxide Interactions
Several decades of research has shown that exposure to elevated CO2 increases
photosynthetic rates (Bernacchi et al.. 2006; Bernacchi et al.. 2005; Tissue et al.. 1999;
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1 Tissue et al.. 1997; Will and Ceulemans. 1997). decreases stomatal conductance
2 (Ainsworth and Rogers. 2007; Paoletti et al.. 2007; Bernacchi et al.. 2006; Leakey et al..
3 2006; Medlyn et al., 2001) and generally increases the growth of plants(McCarthy et al..
4 2009; Norby et al.. 2005). This is in contrast to the decrease in photosynthesis and growth
5 in many plants that are exposed to elevated O3. The interactive effects on vegetation have
6 been the subject of research in the past two decades due to the implications on
7 productivity and water use of ecosystems. This area of research was discussed in detail in
8 AX9.3.8.1 of the 2006 O3 AQCD and the conclusions made then are still relevant (U.S.
9 EPA. 2006b).
10 The bulk of the available evidence shows that, under the various experimental conditions
11 used (which almost exclusively employed abrupt or "step" increases in CO2
12 concentration, as discussed below), increased CO2 levels (ambient + 200 to 400 ppm)
13 may protect plants from the adverse effects of O3 on growth. This protection may be
14 afforded in part by CO2 acting together with O3 in inducing stomatal closure, thereby
15 reducing O3 uptake, and in part by CO2 reducing the negative effects of O3 on Rubisco
16 and its activity in CO2-fixation. Although both CO2-induced and O3-induced decreases in
17 stomatal conductance have been observed primarily in short-term studies, recent data
18 show a long-term and sustained reduction in stomatal conductance under elevated CO2
19 for a number of species (Ainsworth and Long. 2005; Ellsworth et al.. 2004; Gunderson et
20 al.. 2002). Instances of increased stomatal conductance have also been observed in
21 response to O3 exposure, suggesting partial stomatal dysfunction after extended periods
22 of exposure (Paoletti and Grulke. 2010; Grulke et al.. 2007a; Maier-Maercker. 1998).
23 Important caveats must be raised with regard to the findings presented in published
24 research. The first caveat concerns the distinctly different natures of the exposures to O3
25 and CO2 experienced by plants in the field. Changes in the ambient concentrations of
26 these gases have very different dynamics. In the context of climate change, CO2 levels
27 increase relatively slowly (globally 2 ppm/year) and may change little over several
28 seasons of growth. On the other hand, O3 presents a fluctuating stressor with
29 considerable hour-to-hour, day-to-day and regional variability (Polle and Pell. 1999).
30 Almost all of the evidence presented comes from experimentation involving plants
31 subjected to an abrupt step increase to a higher, steady CO2 concentration. In contrast, the
32 O3 exposure concentrations usually varied from day to day. Luo and Reynolds (1999).
33 Hui et al. (2002). and Luo (2001) noted the difficulties in predicting the likely effects of a
34 gradual CO2 increase from experiments involving a step increase or those using a range
35 of CO2 concentrations. It is also important to note that the levels of elevated CO2 in
36 many of the studies will not be experienced in the field for 30 or 40 years, but elevated
37 levels of O3 can occur presently in several areas of the U.S. Therefore, the CO2 * O3
38 interaction studies may be less relevant for current ambient conditions.
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1 Another caveat concerns the interactions of O3 and CO2 with other climatic variables,
2 such as temperature and precipitation. In light of the key role played by temperature in
3 regulating physiological processes and modifying plant response to increased CO2 levels
4 (Morison and Lawlor. 1999; Long. 1991) and the knowledge that relatively modest
5 increases in temperature may lead to dramatic consequences in terms of plant
6 development (Lawlor. 1998). it is important to consider that studying CO2 and O3
7 interactions alone may not create a complete understanding of effects on plants under
8 future climate change.
9.4.9 Insects and Other Wildlife
9.4.9.1 Insects
9 Insects may respond indirectly to changes to plants (i.e., increased reactive oxygen
10 species, altered phytochemistry, altered nutrient content) that occur under elevated O3
11 conditions, or O3 can have a direct effect on insect performance (Menendez et al.. 2009).
12 Effects of O3 on insects occur at the species level (i.e., growth, survival, reproduction,
13 development, feeding behavior) and at the population and community-level (i.e.,
14 population growth rate, community composition). In general, effects of O3 on insects are
15 highly context- and species-specific (Lindroth. 2010; Bidart-Bouzat and Imeh-Nathaniel.
16 2008). Furthermore, plant responses to O3 exposure and herbivore attack have been
17 demonstrated to share signaling pathways, complicating characterization of these
18 stressors (Lindroth. 2010; Menendez et al., 2010. 2009). Although both species-level and
19 population and community-level responses to elevated O3 are observed in field and
20 laboratory studies discussed below, there is no consensus on how insects respond to
21 feeding on O3-exposed plants.
Species-Level Responses
22 In considering insect growth, survival and reproduction in elevated O3 conditions, several
23 studies have indicated an effect while others have found no correlation. The performance
24 of five herbivore species (three moths and two weevils) was assessed in an OTC
25 experiment at 2 x ambient concentration (Peltonen et al.. 2010). Growth of larvae of the
26 Autumnal moth, Epirrita autumna, was significantly decreased in the O3 treatment while
27 no effects were observed in the other species. In an aphid oviposition preference study
28 using birch buds grown in a three year OTC experiment, O3 had neither a stimulatory or
29 deterring effect on egg-laying (Peltonen et al.. 2006). Furthermore, changes in birch bud
30 phenolic compounds associated with the doubled ambient concentrations of O3 did not
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1 correlate with changes in aphid oviposition (Peltonen et al., 2006). Reproduction in
2 Popilliajaponica, that were fed soybeans and grown under elevated O3 appeared to be
3 unaffected (O'Neill et al., 2008). In a meta-analysis of effects of elevated O3 on 22
4 species of trees and 10 species of insects, the rates of survival, reproduction and food
5 consumption were typically unaffected while development times were reduced and pupal
6 masses were increased (Valkama et al.. 2007).
7 At the Aspen FACE site insect performance under elevated (50-60 ppb) O3 conditions
8 (approximately 1.5 x background ambient levels of 30-40 ppb O3) have been considered
9 for several species. Cumulative fecundity of aphids (Cepegillettea betulaefoliae), that
10 were reared on O3-exposed paper birch (Betula papyrifera) trees, was lower than aphids
11 from control plots (Awmack et al.. 2004). No effects on growth, development, adult
12 weight, embryo number and birth weight of newborn nymphs were observed. In a study
13 conducted using three aspen genotypes, performance of the aspen beetle (Chrysomela
14 crochi) decreased across all parameters measured (development time, adult mass and
15 survivorship) under elevated O3 (Vigue and Lindroth. 2010). There was an increase in the
16 development time of male and female aspen beetle larvae although the percentages varied
17 across genotypes. Decreased beetle adult mass and survivorship was observed across all
18 genotypes under elevated O3 conditions. Another study from the Aspen FACE site, did
19 not find any significant effects of elevated O3 on performance (longevity, fecundity,
20 abundance) of the invasive weevil (Polydrusus senceus) (Hillstrom et al.. 201 Ob).
21 Since the 2006 O3 AQCD, several studies have considered the effect of elevated O3 on
22 feeding behavior of insects. In a feeding preference study, the common leaf weevil
23 (Phyllobius pyri) consumed significantly more leaf discs from one aspen clone when
24 compared to a second clone under ambient air conditions (Freiwald et al., 2008). In a
25 moderately elevated O3 environment (1.5 x ambient), this preference for a certain aspen
26 clone was less evident, however, leaves from O3-exposed trees were significantly
27 preferred to leaves grown under ambient conditions. Soybeans grown under enriched O3
28 had significantly less loss of leaf tissue to herbivory in August compared to earlier in the
29 growing season (July) when herbivory was not affected (Hamilton et al.. 2005). Other
30 plant-herbivore interactions have shown no effects of elevated O3 on feeding. Feeding
31 behavior of Japanese beetles (P. japonicd) appeared to be unchanged when beetles were
32 fed soybean leaves grown under elevated O3 conditions (O'Neill et al.. 2008). At the
33 Aspen FACE site, feeding by the invasive weevil (Polydrusus senceus), as measured by
34 leaf area consumption, was not significantly different between foliage that was grown
35 under elevated O3 versus ambient conditions (Hillstrom et al.. 2010b).
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Population-Level and Community-Level Responses
1 Recent data on insects provide evidence of population-level and community-level
2 responses to O3. Elevated levels of O3 can affect plant phytochemistry and nutrient
3 content which in turn can alter population density and structure of the associated
4 herbivorous insect communities and impact ecosystem processes (Cornelissen. 2011;
5 Lindroth. 2010). In 72-hour exposures to elevated O3, mean relative growth rate of the
6 aphid Diuraphis noxia increased with ozone concentration suggesting that more rapid
7 population growth may occur when atmospheric O3 is elevated (Summers et al., 1994,
8 735955). In a long-term study of elevated O3 on herbivore performance at the Aspen
9 FACE site, individual performance and population-level effects of the aphid
10 C. betulaefoliae were assessed. Elevated O3 levels had a strong positive effect on the
11 population growth rates of the aphids; although effects were not detected by measuring
12 growth, development, adult weight, embryo number or birth weight of newborn nymphs
13 (Awmack et al.. 2004). Conversely, a lower rate of population growth was observed in
14 aphids previously exposed to O3 in an OTC (Menendez et al., 2010). No direct effects of
15 O3 were observed; however, nymphs born from adults exposed to and feeding on O3
16 exposed plants were less capable of infesting new plants when compared to nymphs in
17 the control plots (Menendez et al.. 2010). Elevated O3 reduced total arthropod abundance
18 by 17% at Aspen FACE, largely as a result of the negative effects on parasitoids,
19 although phloem-feeding insects may benefit (Hillstrom and Lindroth. 2008). Herbivore
20 communities affected by O3 and N were sampled along an air pollution gradient in the
21 Los Angeles basin (Jones and Paine. 2006). Abundance, diversity, and richness of
22 herbivores were not affected. However, a shift in community structure, from phloem-
23 feeding to chewing dominated communities, was observed along the gradient. No
24 consistent effect of elevated O3 on herbivory or insect population size was detected at
25 SoyFACE (O'Neill etal.. 2010: Dermodv et al.. 2008).
26 Evidence of modification of insect populations and communities in response to elevated
27 O3 includes genotypic and phenotypic changes. In a study conducted at the Aspen FACE
28 site, elevated O3 altered the genotype frequencies of the pea aphid (Acyrthosiphon pi sum)
29 grown on red clover (Trifoliumpratense) over multiple generations (Mondor et al..
30 2005). Aphid color was used to distinguish between the two genotypes. Ozone increased
31 the genotypic frequencies of pink-morph:green-morph aphids from 2:1 to 9:1, and
32 depressed wing-induction responses more strongly in the pink than the green genotype
33 (Mondor et al.. 2005). Growth and development of individual green and pink aphids
34 reared as a single genotype or mixed genotypes were unaffected by elevated O3 (Mondor
35 et al.. 2010). However, growth of pea aphid populations is not readily predictable using
36 individual growth rates.
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9.4.9.2 Wildlife
Herpetofauna
1 Since the 2006 O3 AQCD, direct effects of O3 exposure including physiological changes
2 and alterations of ecologically important behaviors such as feeding and thermoregulation
3 have been observed in wildlife. These studies have been conducted in limited laboratory
4 exposures, and the levels of O3 treatment (e.g. 0.2-0.8 ppm) were often unrealistically
5 higher than the ambient levels. Amphibians may be especially vulnerable to airborne
6 oxidants due to the significant gas exchange that occurs across the skin (Andrews et al..
7 2008; Dohm et al., 2008). Exposure to 0.2 ppm to 0.8 ppm O3 for 4 hours resulted in a
8 decrease of oxygen consumption and depressed lung ventilation in the California tree
9 frog Pseudacris cadaverina (Mautz and Dohm. 2004). Following a single 4-h exposure to
10 O3, reduced pulmonary macrophage phagocytosis was observed at 1 and 24 hours post
11 exposure in the marine toad (Bufo marinus) indicating an effect on immune system
12 function (Dohm et al.. 2005). There was no difference in macrophage function at
13 48 hours post exposure in exposed and control individuals.
14 Behavioral effects of O3 observed in amphibians include responses to minimize the
15 surface area of the body exposed to the air and a decrease in feeding rates (Dohm et al..
16 2008; Mautz and Dohm. 2004). The adoption of a low-profile "water conservation
17 posture" during O3 exposure was observed in experiments with the California tree frog
18 (Mautz and Dohm. 2004). Marine toads, Bufo marinus, exposed to 0.06 (iL/L O3 for
19 4 hours ate significantly fewer mealworms at 1 hour and 48 hours post exposure than
20 control toads (Dohm et al.. 2008). In the same study, escape/exploratory behavior as
21 measured by total distance moved was not adversely affected in the O3-exposed
22 individuals as compared to the controls (Dohmet al.. 2008).
23 Water balance and thermal preference in herpetofauna are altered with elevated O3.
24 Marine toads exposed to 0.8 ppm O3 for 4 hours exhibited behavioral hypothermia when
25 temperature selection in the toads was assessed at I, 24 and 48 hours post exposure
26 (Dohm et al.. 2001). Ozone-exposed individuals lost almost 5g more body mass on
27 average than controls due to evaporative water loss. At 24 hours after exposure, the
28 individuals that had lost significant body mass selected lower body temperatures(Dohm
29 et al.. 2001). Behavioral hypothermia was also observed in reptiles following 4-h
30 exposures to 0.6 ppm O3. Exposure of the Western Fence Lizard (Sceloporus
31 occidentalis) at 25°C induced behavioral hypothermia that recovered to control
32 temperatures by 24 hours (Mautz and Dohm. 2004). The behavioral hypothermic
33 response persisted in lizards exposed to O3 at 35°C at 24 hours post exposure resulting in
34 a mean body temperature of 3.3°C over controls.
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Soil Fauna Communities
1 Ozone has also been shown to alter soil fauna communities (Meehan et al.. 2010;
2 Kasurinen et al., 2007; Loranger et al.. 2004). Abundance of Acari (mites and ticks)
3 decreased by 47% under elevated O3 at Aspen FACE site, probably due to the higher
4 secondary metabolites and lower N concentrations in litter and foliage under elevated O3
5 (Loranger et al.. 2004). In another study from the Aspen FACE site, leaf litter collected
6 from aspen grown under elevated O3 conditions were higher in fiber and lignin
7 concentrations than trees grown under ambient conditions. These chemical characteristics
8 of the leaves were associated with increased springtail population growth following
9 10 weeks in a laboratory microcosm (Meehan et al.. 2010). Consumption rates of
10 earthworms fed on leaf litter for 6 weeks from trees grown under elevated O3 conditions
11 and ambient air did not vary significantly between treatments (Meehan et al.. 2010). In
12 another study on juvenile earthworms Lumbricus terrestris, individual growth was
13 reduced when worms were fed high-O3 birch litter from trees exposed for three years to
14 elevated O3 in an OTC system (Kasurinen et al.. 2007). In the same study no significant
15 growth or mortality effects were observed in isopods.
9.4.9.3 Indirect Effects on Wildlife
16 In addition to the direct effects of O3 exposure on physiological and behavioral endpoints
17 observed in the laboratory, there are indirect effects to wildlife. These effects include
18 changes in biomass and nutritive quality of O3-exposed plants (reviewed in Section 9.4.4)
19 that are consumed by wildlife. Reduced digestibility of O3-exposed plants may alter
20 dietary intake and foraging strategies in herbivores. In a study using native highbush
21 blackberry (Rubus argutus) relative feed value of the plants decreased in bushes exposed
22 to double ambient concentrations of O3 (Ditchkoff et al.. 2009). Indirect effects of
23 elevated O3 on wildlife include changes in chemical signaling important in ecological
24 interactions reviewed below.
Chemical Signaling in Ecological Interactions
25 Ozone has been shown to degrade or alter biogenic VOC signals important to ecological
26 interactions including; (1) attraction of pollinators and seed dispersers; (2) defense
27 against herbivory; and (3) predator-prey interactions (Pinto et al.. 2010; McFrederick et
28 al.. 2009; Yuan et al.. 2009; Pinto et al.. 2007a; Pinto et al.. 2007b). Each signal released
29 by emitters has an atmospheric lifetime and a unique chemical signature comprised of
30 different ratios of individual hydrocarbons that is susceptible to atmospheric oxidants
31 such as O3 (Yuan et al.. 2009; Wright et al.. 2005). Under elevated O3 conditions, these
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1 olfactory cues may travel shorter distances before losing their specificity (McFrederick et
2 al.. 2009; McFrederick et al.. 2008). Additional non-phytogenic VOC-mediated
3 interrelationships with the potential to be modified by O3 include territorial marking,
4 pheromones for attraction of mates and various social interactions including scent trails,
5 nestmate recognition and signals involved in aggregation behaviors (McFrederick et al..
6 2009). For example, the alcohols, ketones and aldehydes comprising sex pheromones in
7 moths could be especially vulnerable to degradation by O3, since some males travel >100
8 m to find mates (Carde and Haynes. 2004). In general, effects of O3 on scent-mediated
9 ecological interactions are highly context- and species-specific (Lindroth. 2010; Bidart-
10 Bouzat and Imeh-Nathaniel. 2008).
Pollination and Seed Dispersal
11 Phytogenic VOC's attract pollinators and seed dispersers to flowers and fruits (Dudareva
12 et al., 2006; Theis and Raguso. 2005). These floral scent trails in plant-insect interactions
13 may be destroyed or transformed by O3 (McFrederick et al.. 2008). Using a Lagrangian
14 model, the rate of destruction of phytogenic VOC's was estimated in air parcels at
15 increasing distance from a source in response to increased regional levels of O3, hydroxyl
16 and nitrate radicals (McFrederick et al., 2008). Based on the model, the ability of
17 pollinators to locate highly reactive VOCs from emitting flowers may have decreased
18 from kilometers during pre-industrial times to <200 m at current ambient conditions
19 (McFrederick et al.. 2008). Scents that travel shorter distances (0-10 m) are less
20 susceptible to air pollutants, while highly reactive scents that travel longer distances (10
21 to 100's of meters), are at a higher risk for degradation (McFrederick et al.. 2009). For
22 example, male euglossine bees can detect bait stations from a distance of at least one
23 kilometer (Dobson. 1994).
Defense Against Herbivory
24 Ozone can alter the chemical signature of VOCs emitted by plants and these VOCs are
25 subsequently detected by herbivores (Blande et al.. 2010; Iriti and Faoro. 2009; Pinto et
26 al.. 2007a; Vuorinen et al.. 2004; Jackson et al.. 1999; Cannon. 1990). These
27 modifications can make the plant either more attractive or repellant to phytophagous
28 insects (Pinto etal.. 2010). For example, under elevated O3, the host plant preference by
29 forest tent caterpillars increased for birch compared to aspen (Agrell et al., 2005). Ozone-
30 induced emissions from red spruce needles were found to repel spruce budworm larvae
31 (Cannon. 1990). Transcriptional profiles of field grown soybean (Glycine max) grown in
32 elevated O3 conditions were altered due to herbivory by Japanese beetles. The herbivory
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1 resulted in a higher number of transcripts in the leaves of O3-exposed plants and up-
2 regulation of antioxidant metabolism associated with plant defense (Casteel et al.. 2008).
3 Ozone may modify signals involved in plant-to-plant interactions and plant defense
4 against pathogens (Blandeetal., 2010; Pinto etal., 2010; McFrederick et al., 2009; Yuan
5 et al.. 2009). In a recent study with lima beans, 80 ppb O3 degraded several herbivore-
6 induced VOCs, reducing the distance over which plant-to-plant signaling occurred
7 (Blandeetal.,2010).
Predator-Prey Interactions
8 Elevated O3 conditions are associated with disruption of pheromone-mediated
9 interactions at higher trophic levels (e.g., predators and parasitoids of herbivores). In a
10 study from the Aspen FACE site, predator escape behaviors of the aphid (Chatophorus
11 stevensis) were enhanced on O3-fumigated aspen trees although the mechanism of this
12 response remains unknown (Tvlondor et al.. 2004). The predatory mite Phytoseiulus
13 persimilis can distinguish between the VOC signature of ozonated lima bean plants and
14 ozonated lima bean plants simultaneously damaged by T. urticae (Vuorinen et al.. 2004)
15 however, other tritrophic interactions have shown no effect (Pinto et al.. 2007b).
16 There are few studies that consider host location behaviors of parasites under elevated
17 O3. In closed chambers fumigated with O3, the searching efficiency and proportion of the
18 host larval fruit flies parasitized by Asobara tabida, declined when compared to filtered
19 air controls (Gate etal.. 1995). The host location behavior and rate of parasitism of the
20 wasp (Coesiaplutellae) on Plutella xylostella-mfested potted cabbage plants was tested
21 under ambient and doubled O3 conditions in an open-air fumigation system (Pinto et al..
22 2008). The number of wasps found in the field and the percentages of parasitized larvae
23 were not significantly different from controls under elevated O3.
24 Elevated O3 has the potential to perturb specialized food-web communication in
25 transgenic crops. In insect-resistant oilseed rape Brassica napus grown under 100 ppb O3
26 in a growth chamber, reduced feeding damage by Putella xylostella led to deceased
27 attraction of the endoparasitoid (Costesia vestalis), however this tritrophic interaction
28 was influenced by the degree of herbivore feeding (Himanen et al.. 2009a; Himanen et
29 al.. 2009b). Under chronic O3-exposure, the insect resistance trait BT crylAc in
30 transgenic B. napus was higher than the control (Himanen et al., 2009c). There was a
31 negative relative growth rate of the Bt target herbivore, P. xylostella, in all O3 treatments.
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9.4.9.4 Summary
1 New information on O3 effects on insects and other wildlife is limited to a few species
2 and there is no consensus on how these organisms respond to elevated O3 Studies
3 published since the last review show impacts of elevated O3 on both species-level
4 responses (reproduction, growth, feeding behavior) and community and ecosystem-level
5 responses (population growth, abundance, shift in community structure) in some insects
6 and soil fauna. Changes in ecologically important behaviors such as feeding and
7 thermoregulation have recently been observed with O3 exposure in amphibians and
8 reptiles, however, these responses occur at concentrations of O3 much higher that
9 ambient levels.
10 New information available since the last review considers the effects of O3 on chemical
11 signaling in insect and wildlife interactions. Specifically, studies on O3 effects on
12 pollination and seed dispersal, defenses against herbivory and predator-prey interactions
13 all consider the ability of O3 to alter the chemical signature of VOCs emitted during these
14 pheromone-mediated events. The effects of O3 on chemical signaling between plants,
15 herbivores and pollinators as well as interactions between multiple trophic levels is an
16 emerging area of study that may result in further elucidation of O3 effects at the species,
17 community and ecosystem-level.
9.5 Effects-Based Air Quality Exposure Indices and Dose
Modeling
9.5.1 Introduction
18 Exposure indices are metrics that quantify exposure as it relates to measured plant
19 damage (e.g., reduced growth). They are summary measures of monitored ambient O3
20 concentrations over time, intended to provide a consistent metric for reviewing and
21 comparing exposure-response effects obtained from various studies. Such indices may
22 also provide a basis for developing a biologically-relevant air quality standard for
23 protecting vegetation and ecosystems. Effects on plant growth and/or yield have been a
24 major focus of the characterization of O3 impacts on plants for purposes of the air quality
25 standard setting process (U.S. EPA. 2007b. 1996e. 1986). The relationship of O3 and
26 plant responses can be characterized quantitatively as "dose-response" or "exposure-
27 response." The distinction is in how the pollutant concentration is expressed: "dose" is
28 the pollutant concentration absorbed by the leaf over some time period, and is very
29 difficult to measure directly, whereas "exposure" is the ambient air concentration
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1 measured near the plant over some time period, and summarized for that period using an
2 index. Exposure indices have been most useful in considering the form of secondary O3
3 NAAQS, in large part because they only require ambient air quality data rather than more
4 complex indirect calculations of dose to the plant. The attributes of exposure indices that
5 are most relevant to plant damage are the weighting of O3 concentrations and the daily
6 and seasonal time-periods. Several different types of exposure indices are discussed in
7 Section 9.5.2.
8 Form a theoretical perspective, a measure of plant O3 uptake or dose from ambient air
9 (either rate of uptake or cumulative seasonal uptake) might be a belter predictor of O3
10 damage to plants than an exposure index and may be useful in improving risk assessment.
11 An uptake estimate would have to integrate all those environmental factors that influence
12 stomatal conductance, including but not limited to temperature, humidity, and soil water
13 status (Section 9.5.4). Therefore, uptake values are generally obtained with simulation
14 models that require knowledge of species- and site-specific values for the variables
15 mentioned. However, a limitation of modeling dose is that environmental variables are
16 poorly characterized. In addition, it has also been recognized that O3 detoxification
17 processes and the temporal dynamics of detoxification must be taken into account in dose
18 modeling (Heath et al.. 2009) (Section 9.5.4). Because of this, research has focused
19 historically on predictors of O3 damage to plants based only on exposure as a summary
20 measure of monitored ambient pollutant concentration over some integral of time, rather
21 than dose (U.S. EPA. 1996c: Costa et al.. 1992; Leeetal.. 1988b: U.S. EPA. 1986;
22 Lefohn and Benedict. 1982: O'Gara. 1922).
9.5.2 Description of Exposure Indices Available in the Literature
23 Mathematical approaches for summarizing ambient air quality information in biologically
24 meaningful forms for O3 vegetation effects assessment purposes have been explored for
25 more than 80 years (U.S. EPA. 1996b: O'Gara. 1922). In the context of national standards
26 that protect for "known or anticipated" effects on many plant species in a variety of
27 habitats, exposure indices provide a numerical summary of very large numbers of
28 ambient observations of concentration over extended periods. Like any summary statistic,
29 exposure indices retain information on some, but not all, characteristics of the original
30 observations. Several indices have been developed to attempt to incorporate some of the
31 biological, environmental, and exposure factors that influence the magnitude of the
32 biological response and contribute to observed variability (Hogsett et al.. 1988). In the
33 1996 O3 AQCD, the exposure indices were arranged into five categories; (1) One event,
34 (2) Mean, (3) Cumulative, (4) Concentration weighted, and (5) Multicomponent, and
35 were discussed in detail (Lee etal.. 1989). Figure 9-9 illustrates how several of the
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1
2
3
4
5
6
indices weight concentration and accumulate exposure. For example, the SUM06 index
(panel a) is a threshold-based approach wherein concentrations below 0.06 ppm are given
a weight of zero and concentrations above 0.06 ppm are given a weight of 1.0 that is
summed usually over 3 to 6 months . The Sigmoid approach (panel b), which is similar to
the W126 index, is a non-threshold approach wherein all concentrations are given a
weight that increases from zero to 1.0 with increasing concentration and summed.
0.15
Ppm 2ndHDM->
M-7 = 0.05 ppm
) 2 4 6 8
Day
1
0.10
0.05
0.00
Source: Used with permission from Air and Waste Management Association (Tingevet al.. 1991)
(a) SUM06: the upper graphic illustrates an episodic exposure profile; the shaded area under some of the peaks illustrates the
concentrations greater than or equal to 0.06 ppm that are accumulated in the index. The insert shows the concentration weighting (0
to 1) function. The lower portion of the graphic illustrates how concentration is accumulated over the exposure period, (b) SIGMOID:
the upper graphic illustrates an episodic exposure profile; the variable shaded area under the peaks illustrates the concentration-
dependent weights that are accumulated in the index. The insert shows the sigmoid concentration weighting function. This is similar
to the W126 function. The lower portion of the graphic illustrates how concentration is accumulated over the exposure period, (c)
second HDM and M-7: the upper graphic illustrates an episodic exposure profile. The lower portion of the graphic illustrates that the
second HDM considers only a single exposure peak, while the M-7 (average of 7-h daily means) applies a constant exposure value
over the exposure period.
Figure 9-9 Diagrammatic representation of several exposure indices
illustrating how they weight concentration and accumulate
exposure.
Various factors with known or suspected bearing on the exposure-response relationship,
including concentration, time of day, respite time, frequency of peak occurrence, plant
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1 phenology, predisposition, etc., have been weighted with various functions in a large set
2 of indices. The resulting indices were evaluated by ranking them according to the
3 goodness-of-fit of a regression model of growth or yield response (Lee etal. 1989). The
4 statistical evaluations for each of these indices were completed using growth or yield
5 response data from many earlier exposure studies (e.g., NCLAN). This retrospective
6 approach was necessary because there were no studies specifically designed to test the
7 goodness of fit of the various indices. The goodness of fit of a set of linear and nonlinear
8 models for exposure-response was ranked as various proposed indices were used in turn
9 to quantify exposure. This approach provided evidence for the best indices. The results of
10 retrospective analyses are described below.
11 Most of the early retrospective studies reporting regression approaches used data from the
12 NCLAN program or data from Corvallis, Oregon or California (Costa etal.. 1992; Lee et
13 al.. 1988b: Lefohn et al.. 1988: Musselman et al.. 1988: Lee et al.. 1987: U.S. EPA.
14 1986). These studies were previously reviewed by the EPA (U.S. EPA, 1996c; Costa et
15 al.. 1992) and were in general agreement that the best fit to the data resulted from using
16 cumulative concentration-weighted exposure indices (e.g. W126, SUM06). Lee et al.
17 (1987) suggested that exposure indices that included all the 24-h data performed better
18 than those that used only 7 hours of data; this was consistent with the conclusions of
19 Heagle et al. (1987) that plants receiving exposures for an additional 5-h/day showed
20 10% greater yield loss than those exposed for 7-h/day. In an analysis using the National
21 Crop Loss Assessment Network (NCLAN) data, Lee et al. (1988) found several indices
22 which only cumulated and weighted higher concentrations (e.g., W126, SUM06, SUM08,
23 and AOT40) performed very well. Amongst this group no index had consistently better
24 fits than the other indices across all studies and species (Heagle et al.. 1994b; Lefohn et
25 al.. 1988; Musselman et al.. 1988). Lee et al. (1988) found that adding phenology
26 weighting to the index somewhat improved the performance of the indices. The "best"
27 exposure index was a phenologically weighted cumulative index, with sigmoid weighting
28 on concentration and a gamma weighting function as a surrogate for plant growth stage.
29 This index provided the best statistical fit when used in the models under consideration,
30 but it required data on species and site conditions, making specification of weighting
31 functions difficult for general use.
32 Other factors, including predisposition time (Hogsett et al.. 1988; McCool et al.. 1988)
33 and crop development stage (Tingey et al.. 2002; Heagle et al.. 1991) contributed to
34 variation in the biological response and suggested the need for weighting O3
35 concentrations to account for predisposition time and phenology. However, the roles of
36 predisposition and phenology in plant response vary considerably with species and
37 environmental conditions; therefore, specification of a weighting function for general use
38 in characterizing plant exposure has not been possible.
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1 European scientists took a similar approach in developing indices describing growth and
2 yield loss in crops and tree seedlings, using OTCs with modified ambient exposures, but
3 many fewer species and study locations were employed in the European studies. There is
4 evidence from some European studies that a lower (Pleijel et al.. 1997) or higher (Finnan
5 et al., 1997; Finnan et al., 1996) cutoff value in indices with a threshold may provide a
6 better statistical fit to the experimental data. Finnan et al. (1997) used seven exposure
7 studies of spring wheat to confirm that cumulative exposure indices emphasizing higher
8 O3 concentrations were best related to plant response and that cumulative exposure
9 indices using weighting functions, including cutoff concentrations, allometric and
10 sigmoidal, provided a better fit and that the ranking of these indices differed depending
11 on the exposure-response model used. Weighting those concentrations associated with
12 sunshine hours in an attempt to incorporate an element of plant uptake did not improve
13 the index performance (Finnan et al.. 1997). A more recent study using data from several
14 European studies of Norway spruce, analyzed the relationship between relative biomass
15 accumulation and several cumulative, weighted indices, including the AOT40 (area over
16 a threshold of 40ppb) and the SUM06 (Skarby et al.. 2004). All the indices performed
17 relatively well in regressing biomass and exposure index, with the AOT20 and AOT30
18 doing slightly better than others (r2 = 0.46-0.47). In another comparative study of four
19 independent data sets of potato yield and different cumulative uptake indices with
20 different cutoff values, a similarly narrow range of r2 was observed (r2 = 0.3-0.4) (Pleijel
21 et al.. 2004b).
22 In Europe, the cutoff concentration-weighted index AOT40 was selected in developing
23 exposure-response relationships based on OTC studies of a limited number of crops and
24 trees (Grunhage and Jager. 2003). The United Nations Economic Commission for Europe
25 (UNECE. 1988) adopted the critical levels approach for assessment of O3 risk to
26 vegetation across Europe. As used by the UNECE, the critical levels are not like the air
27 quality regulatory standards used in the U.S., but rather function as planning targets for
28 reductions in pollutant emissions to protect ecological resources. Critical levels for O3 are
29 intended to prevent long-term deleterious effects on the most sensitive plant species
30 under the most sensitive environmental conditions, but not intended to quantify O3
31 effects. A critical level was defined as "the concentration of pollutant in the atmosphere
32 above which direct adverse effects on receptors, such as plants, ecosystems, or materials
33 may occur according to present knowledge" (UNECE. 1988). The nature of the "adverse
34 effects" was not specified in the original definition, which provided for different levels
35 for different types of harmful effect (e.g., visible injury or loss of crop yield). There are
36 also different critical levels for crops, forests, and semi-natural vegetation. The caveat,
37 "according to present knowledge" is important because critical levels are not rigid; they
38 are revised periodically as new scientific information becomes available. For example,
39 the original critical level for O3 specified concentrations for three averaging times, but
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1 further research and debate led to the current critical level being stated as the cumulative
2 exposure (concentration x hours) over a cutoff concentration of 40 ppb (AOT40) (Tuhrer
3 etal.. 1997).
4 More recently in Europe, a decision was made to work towards a flux-based approach
5 (see section 9.5.4) for the critical levels ("Level II"), with the goal of modeling O3 flux-
6 effect relationships for three vegetation types: crops, forests, and semi-natural vegetation
7 (Grunhage and Jager. 2003). Progress has been made in modeling flux (U.S. EPA. 2006b)
8 and the Mapping Manual is being revised (Ashmore et al.. 2004a. b; Grennfelt. 2004;
9 Karlsson et al.. 2003). The revisions may include a flux-based approach for three crops:
10 wheat, potatoes, and cotton. However, because of a lack of flux-response data, a
11 cumulative, cutoff concentration-based (AOTx) exposure index will remain in use for the
12 near future for most crops and for forests and semi-natural herbaceous vegetation
13 (Ashmore et al.. 2004b)
14 In both the U.S. and Europe, the adequacy of these numerical summaries of exposure in
15 relating biomass and yield changes have, for the most part, all been evaluated using data
16 from studies not necessarily designed to compare one index to another (Skarby et al..
17 2004; Lee etal.. 1989; Lefohnetal.. 1988). Very few studies in the U.S. have addressed
18 this issue since the 2006 O3 AQCD. McLaughlin et al. (2007a) reported that the
19 cumulative exposure index of AOT60 related well to reductions in growth rates at forest
20 sites in the southern Appalachian Mountains. However, the authors did not report an
21 analysis to compare multiple indices. Overall, given the available data from previous O3
22 AQCDs and the few recent studies, the cumulative, concentration-weighted indices
23 perform better than the peak or mean indices. It is still not possible, however, to
24 distinguish the differences in performance among the cumulative, concentration-weighted
25 indices.
26 The main conclusions from the 1996 and 2006 O3 AQCDs regarding an index based on
27 ambient exposure are still valid. No information has come forth since the 2006 O3 AQCD
28 to alter those conclusions significantly. These key conclusions can be restated as follows:
29 • O3 effects in plants are cumulative;
30 • higher O3 concentrations appear to be more important than lower
31 concentrations in eliciting a response;
32 • plant sensitivity to O3 varies with time of day and plant development stage;
33 and
34 • exposure indices that accumulate the O3 hourly concentrations and
35 preferentially weight the higher concentrations have better statistical fits to
36 growth/yield response than do the mean and peak indices.
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1 Following the 2006 criteria review process (U.S. EPA. 2006b). the EPA proposed an
2 alternative form of the secondary NAAQS for O3 using a cumulative, concentration-
3 weighted exposure index to protect vegetation from damage (72 FR37818). The EPA
4 considered two specific concentration-weighted indices: the cutoff concentration
5 weighted SUM06 and the sigmoid-weighted W126 exposure index (U.S. EPA. 2007b).
6 These two indices performed equally well in predicting the exposure-response
7 relationships observed in the crop and tree seedlings studies (Lee etal.. 1989). At a
8 workshop convened to consider the science supporting these indices (Heck and Cowling.
9 1997) there was a consensus that these cumulative concentration-weighted indices being
10 considered were equally capable of predicting plant response. Below are the definitions
11 of the two cumulative index forms considered in the previous staff paper review (U.S.
12 EPA. 2007b):
13 • SUM06: Sum of all hourly O3 concentrations greater than or equal to
14 0.06 ppm observed during a specified daily and seasonal time window (Figure
15 9-9a).
16 • W126: Sigmoidally weighted sum of all hourly O3 concentrations observed
17 during a specified daily and seasonal time window (Similar to Figure 9-9b).
18 The sigmoidal weighting of hourly O3 concentration is given in the equation
19 below, where C is the hourly O3 concentration in ppm:
1
W =
c 4403e~126C
Equation 9-1
20 The SUM06 and W126 indices have a variety of relevant time windows that may be
21 applied and are discussed in Section 9.5.3.
22 It should be noted that there are some important differences between the SUM06 and
23 W126. When considering the response of vegetation to ozone exposures represented by
24 the threshold (e.g., SUM06) and non-threshold (e.g., W126) indices, the W126 metric
25 does not have a cut-off in the weighting scheme as does SUM06 and thus it includes
26 consideration of potentially damaging exposures below 60 ppb. The W126 metric also
27 adds increasing weight to hourly concentrations from about 40 ppb to about 100 ppb.
28 This is unlike cut-off metrics such as the SUM06 where all concentrations above 60 ppb
29 are treated equally. This is an important feature of the W126 since as hourly
30 concentrations become higher, they become increasingly likely to overwhelm plant
31 defenses and are known to be more detrimental to vegetation (See Section 9.5.3.1).
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9.5.3 Important Components of Exposure Indices
1 In the previous O3 AQCDs it was established that higher hourly concentrations have
2 greater effects on vegetation than lower concentrations (U.S. EPA. 2006b. 1996c).
3 Further, it was determined that the diurnal and seasonal duration of exposure is important
4 for plant response. Weighting of hourly concentrations and the diurnal and seasonal time
5 window of exposure are the most important variables in a cumulative exposure index and
6 will be discussed below. However, these variables must be taken in the context of plant
7 phenology, diurnal conductance rates, plant canopy structure, and detoxification
8 mechanisms of vegetation as well as the climate and meteorology, all of which are
9 determinants of plant response. These more specific factors will be discussed in the
10 uptake and dose modeling section 9.5.4.
9.5.3.1 Role of Concentration
11 The significant role of peak O3 concentrations was established based on several
12 experimental studies (U.S. EPA. 1996c). Several studies (Oksanen and Holopainen.
13 2001; Yun and Laurence. 1999; Nussbaum et al., 1995) have added support for the
14 important role that peak concentrations, as well as the pattern of occurrence, plays in
15 plant response to O3. Oksanen and Holopainen (2001) found that the peak concentrations
16 and the shape of the O3 exposure (i.e., duration of the event) were important determinants
17 of foliar injury in European white birch saplings, but growth reductions were found to be
18 more related to total cumulative exposure. Based on air quality data from 10 U.S. cities,
19 three 4-week exposure treatments having the same SUM06 value were constructed by
20 Yun and Laurence (1999). The authors used different exposure regimes to explore effects
21 of treatments with variable versus uniform peak occurrence during the exposure period.
22 The authors reported that the variable peak exposures were important in causing injury,
23 and that the different exposure treatments, although having the same SUM06, resulted in
24 very different patterns of foliar injury. Nussbaum et al. (1995) also found peak
25 concentrations and the pattern of occurrence to be critical in determining the measured
26 response. The authors recommended that to describe the effect on total forage yield, peak
27 concentrations >0.11 ppm must be emphasized by using an AOT with higher threshold
28 concentrations.
29 A greater role for peak concentrations in effects on plant growth might be inferred based
30 on air quality analyses for the southern California area (Tingey et al.. 2004; Lee et al..
31 2003a). In the late 1960s and 1970s, extremely high O3 concentrations had impacted the
32 San Bernardino National Forest. However, over the past 20+ years, significant reductions
33 in O3 exposure have occurred (Bvtnerowicz et al., 2008; Lee et al., 2003a; Lefohn and
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1 Shadwick. 2000; Davidson. 1993). An illustration of this improvement in air quality is
2 shown by the 37-year history of O3 air quality at the Crestline site in the San Bernardino
3 Mountains (Figure 9-10) (Lee et al., 2003a). Ozone exposure increased from 1963 to
4 1979 concurrent with increased population and vehicular miles, followed by a decline to
5 the present mirroring decreases in precursor emissions. The pattern in exposure was
6 evident in various exposure indices including the cumulative concentration weighted
7 (SUM06), as well as maximum peak event (1-h peak), and the number of days having
8 hourly averaged O3 concentrations greater than or equal to 95 ppb. The number of days
9 having hourly averaged O3 concentrations greater than or equal to 95 ppb declined
10 significantly from 163 days in 1978 to 103 days in 1997. The changes in ambient O3 air
11 quality for the Crestline site were reflected in the changes in frequency and magnitude of
12 the peak hourly concentration and the duration of exposure (Figure 9-10). Considering
13 the role of exposure patterns in determining response, the seasonal and diurnal patterns in
14 hourly O3 concentration did not vary appreciably from year to year over the 37-year
15 period (Lee et al., 2003a).
16 The potential importance of exposure to peak concentrations comes both from results of
17 measures of tree conditions on established plots and from results of model simulations.
18 Across a broad area of the San Bernardino National Forest, the Forest Pest Management
19 (FPM) method of injury assessment indicated an improvement in crown condition from
20 1974 to 1988; and the area of improvement in injury assessment is coincident with an
21 improvement in O3 air quality (Miller and Rechel. 1999). A more recent analysis of
22 forest changes in the San Bernardino National Forest, using an expanded network of
23 monitoring sites, has verified significant changes in growth, mortality rates, basal area,
24 and species composition throughout the area since 1974 (Arbaugh et al.. 2003). A model
25 simulation of ponderosa pine growth over the 40-year period in the San Bernardino
26 National Forest showed a significant impact of O3 exposure on tree growth and indicates
27 improved growth with reduced O3 concentrations. This area has also experienced
28 elevated N deposition and based on a number of environmental indicators, it appears that
29 this area is experiencing N saturation (Fenn et al., 1996). To account for this potential
30 interaction, the model simulations were conducted under conditions of unlimited soil N.
31 The actual interactions are not known. The improvement in growth over the years was
32 attributed to improved air quality, but no distinction was made regarding the relative role
33 of mid-range and higher hourly concentrations, only that improved growth tracked
34 decreasing SUM06, maximum peak concentration, and number of days of hourly O3
35 >95 ppb (Tingev et al.. 2004). A summary of air quality data from 1980 to 2000 for the
36 San Bernardino National Forest area of the number of "mid-range" hourly concentrations
37 indicated no dramatic changes over this 20-year period, ranging from about 1,500 to
38 2,000 hours per year (Figure 9-11). There was a slow increase in the number of mid-
39 range concentrations from 1980 to 1986, which corresponds to the period after
Draft - Do Not Cite or Quote 9-108 September 2011
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1
2
3
4
5
6
7
implementation of the air quality standard. Another sharper increase was observed in the
late 1990s. This pattern of occurrence of mid-range hourly concentrations suggests a
lesser role for these concentration ranges compared to the higher values in either of the
ground-level tree injury observations of the model simulation of growth over the 40-year
period.
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Source: Used with permission from Elsevier Science Ltd. (Lee et a I.. 2003a).
Annual ROG and NOX emissions data for San Bernardino County were obtained from Alexis et al. (2001 a) and the California Air
Resource Board's emission inventory available at http://www.arb.ca.gov/aqd/aqdpage.htm (Cal/EPA, 2010).
Figure 9-10 Trends in May to September 12-h SUM06, peak 1-h ozone
concentration and number of daily exceedances of 95 ppb for the
Crestline site in 1963 to 1999 in relation to trends in mean daily
maximum temperature for Crestline and daily reactive organic
gases (ROG) and oxides of nitrogen (NOx) for San Bernardino
County.
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9-109
September 2011
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Crestline, San Bernardino, CA
Number of Hours 50 - 89 ppb
060710005
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Figure 9-11 The number of hourly average concentrations between 50 and 89
ppb for the period 1980-2000 for the Crestline, San Bernardino
County, CA, monitoring site.
9.5.3.2 Diurnal and Seasonal Exposure
1
2
3
4
5
6
7
8
9
10
Diurnal Exposure
The diurnal patterns of maximal leaf/needle conductance and occurrence of higher
ambient concentrations can help determine which hours during the day over a season
should be included in an exposure index. Stomatal conductance is species and phenology
dependent and is linked to both diurnal and seasonal meteorological activity as well as to
soil/site conditions (e.g., VPD, soil moisture). Daily patterns of leaf/needle conductance
are often highest in midmorning, whereas higher ambient O3 concentrations generally
occurred in early to late afternoon when stomata were often partially closed and
conductances were lower. Total O3 flux depends on atmospheric and boundary layer
resistances, both of which exhibit variability throughout the day. Experimental studies
with tree species demonstrated the decoupling of ambient O3 exposure, peak occurrence,
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September 2011
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1 and gas exchange, particularly in areas of drought (Panek. 2004). Several studies have
2 suggested that ponderosa pine trees in the southern and northern Sierra Nevada
3 Mountains may not be as susceptible to high O3 concentrations as to lower
4 concentrations, due to reduced needle conductance and O3 uptake during the period when
5 the highest concentrations occur (Panek et al., 2002; Panek and Goldstein, 2001; Bauer et
6 al.. 2000; Arbaugh et al.. 1998). Panek et al. (2002) compared direct O3 flux
7 measurements into a canopy of ponderosa pine and demonstrated a lack of correlation of
8 daily patterns of conductance and O3 occurrence, especially in the late season drought
9 period; the authors concluded that a consideration of climate or season was essential,
10 especially considering the role of soil moisture and conductance/uptake. In contrast,
11 Grulke et al. (2002) reported high conductance when O3 concentrations were high in the
12 same species, but under different growing site conditions. The longer-term biological
13 responses reported by Miller and Rechel (1999) for ponderosa pine in the same region,
14 and the general reduction in recent years in ambient O3 concentrations, suggest that
15 stomatal conductance alone may not be a sufficient indicator of potential vegetation
16 injury or damage. Another consideration for the effect of O3 uptake is the diurnal pattern
17 of detoxification capacity of the plant. The detoxification capacity may not follow the
18 same pattern as stomatal conductance (Heath et al.. 2009).
19 The use of a 12-h (8:00 a.m. to 8:00 p.m.) daylight period for a W126 cumulating
20 exposure was based primarily on evidence that the conditions for uptake of O3 into the
21 plant occur mainly during the daytime hours. In general, plants have the highest stomatal
22 conductance during the daytime and in many areas atmospheric turbulent mixing is
23 greatest during the day as well (Uddling etal.. 2010; U.S. EPA. 2006b). However,
24 notable exceptions to maximum daytime conductance are cacti and other plants with
25 crassulacean acid metabolism (CAM photosynthesis) which only open their stomata at
26 night. This section will focus on plants with C3 and C4 photosynthesis, which generally
27 have maximum stomatal conductance during the daytime.
28 Recent reviews of the literature reported that a large number of species had varying
29 degrees of nocturnal stomatal conductance (Caird et al.. 2007; Dawson et al.. 2007;
30 Musselman and Minnick. 2000). The reason for night-time water loss through stomata is
31 not well understood and is an area of active research (e.g. (Christman et al.. 2009;
32 Howard et al.. 2009) Night-time stomata opening may be enhanced by O3 damage that
33 could result in loss of stomatal control, and less complete closure of stomata, than under
34 low O3 conditions (Caird et al., 2007; Grulke et al., 2007b). In general, the rate of
3 5 stomatal conductance at night is much lower than during the day (Caird et al.. 2007).
36 Atmospheric turbulence at night is also often low, which results in stable boundary layers
37 and unfavorable conditions for O3 uptake into vegetation (Finkelstein et al.. 2000).
38 Nevertheless, nocturnal turbulence does intermittently occur and may result in
Draft - Do Not Cite or Quote 9-111 September 2011
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1 nonnegligible O3 flux into the plants. In addition, plants might be more susceptible to O3
2 exposure at night than during the daytime, because of potentially lower plant defenses
3 (Heath et al., 2009; Loreto and Fares. 2007; Musselman et al., 2006; Musselman and
4 Minnick. 2000). For significant nocturnal stomatal flux and O3 effects to occur, specific
5 conditions must exist. A susceptible plant with nocturnal stomatal conductance and low
6 defenses must be growing in an area with relatively high night-time O3 concentrations
7 and appreciable nocturnal atmospheric turbulence. It is unclear how many areas there are
8 in the U.S. where these conditions occur. It may be possible that these conditions exist in
9 mountainous areas of southern California, front-range of Colorado (Turnipseed et al.,
10 2009) and the Great Smoky Mountains of North Carolina and Tennessee. Tobiessen et al.
11 (1982) found that shade intolerant tree species showed opening of stomata in the dark and
12 did not find this in shade tolerant species. This may indicate shade intolerant trees may be
13 more likely to be susceptible to O3 exposure at night. More information is needed in
14 locations with high night-time O3 to assess the local O3 patterns, micrometeorology and
15 responses of potentially vulnerable plant species.
16 Several field studies have attempted to quantify night-time O3 uptake with a variety of
17 methods. However, many of these studies have not linked the night-time flux to measured
18 effects on plants. Grulke et al. (2004) showed that the stomatal conductance at night for
19 ponderosa pine in the San Bernardino National Forest (CA) ranged from one tenth to one
20 fourth that of maximum daytime stomatal conductance. In June, at a high-elevation site, it
21 was calculated that 11% of the total daily O3 uptake of pole-sized trees occurred at night.
22 In late summer, however, O3 uptake at night was negligible. However, this study did not
23 consider the turbulent conditions at night. Finklestein et al. (2000) investigated O3
24 deposition velocity to forest canopies at three different sites. The authors found the total
25 flux (stomatal and non-stomatal) to the canopy to be very low during night-time hours as
26 compared to day-time hours. However, the authors did note that higher nocturnal
27 deposition velocities at conifer sites may be due to some degree of stomatal opening at
28 night (Finkelstein et al.. 2000). Work by Mereu et al. (2009) in Italy on Mediterranean
29 species indicated that nocturnal uptake was from 10 to 18% of total daily uptake during a
30 weak drought and up to 24% as the drought became more pronounced. The proportion of
31 night-time uptake was greater during the drought due to decreases in daytime stomatal
32 conductance (Mereu et al.. 2009). In a study conducted in California, Fares et al. (Fares et
33 al., 2011) reported that calculated mean percentages of nocturnal uptake were 5%, 12.5%,
34 6.9% of total O3 uptake for lemon, mandarin, and orange, respectively. In another recent
35 study at the Aspen FACE site in Wisconsin, calculated leaf-level stomatal O3 flux was
36 near zero from the night-time hours of 8:00 p.m. to 5:00 a.m. (Uddling et al.. 2010). This
37 was likely due to low horizontal wind speed (>1 m/s) and low O3 concentrations
38 (<25 ppb) during those same night-time hours (Figure 9-12).
Draft - Do Not Cite or Quote 9-112 September 2011
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5 ID 15 20
Time of day
Figure 9-12 Mean diurnal, (a) conductance through boundary layer and
stomata (gbs), (b) Ozone concentration, and leaf-level stomatal
ozone flux without flux cut-off threshold (FstOi) in control plots
from mid-June through August in (c) 2004 and (d) 2005 in the Aspen
FACE experiment. Subscripts "max" and "min" refer to stomatal
fluxes calculated neglecting and accounting for potential non-
stomatal ozone flux, respectively.
1
2
3
4
5
6
1
A few studies have tested the biological effects of night-time O3 exposure on vegetation
in controlled chambers. Biomass of ponderosa pine seedlings was significantly reduced
when seedlings were exposed to either daytime or nighttime episodic profiles (Lee and
Hogsett 1999). However, the biomass reductions were much greater with daytime peak
concentrations than with nighttime peak concentrations. Similarly, birch cuttings grown
in field chambers that were exposed to O3 at night only, daytime only, and 24 hours
showed similar reductions in biomass in night only and day only treatments. Birch
seedling showed greater reductions in growth in 24-h exposures than those exposed to O3
at night or day only (Matyssek et al.. 1995). Field mustard (Brassica rapd) plants
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1 exposed to O3 during the day or night showed little significant difference in the amounts
2 of injury or reduced growth response to O3 treatment, although the stomatal conductance
3 was 70-80% lower at night (Winner et al., 1989). These studies show that effects can be
4 seen with night-time exposures to O3 but when atmospheric conditions are stable at night,
5 it is uncertain how these exposures may affect plants and trees with complex canopies in
6 the field.
Seasonal Exposure
7 Vegetation across the U.S. has widely varying periods of physiological activity during the
8 year due to variability in climate and phenology. In order for a particular plant to be
9 vulnerable to O3 pollution, it must have foliage and be physiologically active. Annual
10 crops are typically grown for periods of two to three months. In contrast, perennial
11 species may be photosynthetically active longer (up to 12 months each year for some
12 species) depending on the species and where it is grown. In general, the period of
13 maximum physiological activity and thus, potential O3 uptake for vegetation coincides
14 with some or all of the intra-annual period defined as the O3 season, which varies on a
15 state-by-state basis (Figure 3-19). This is because the high temperature and high light
16 conditions that typically promote the formation of tropospheric O3 also promote
17 physiological activity in vegetation. There are very limited exceptions to this pattern
18 where O3 can form in the winter in areas in the western U.S. with intense natural gas
19 exploration (Pinto. 2009). but this is typically when plants are dormant and there is little
20 chance of O3 uptake. The selection of any single window of time for a national standard
21 to consider hourly O3 concentrations represents a compromise, given the significant
22 variability in growth patterns and lengths of growing season among the wide range of
23 vegetation species that may experience adverse effects associated with O3 exposure.
24 Various intra-annual averaging and accumulation time periods have been considered for
25 the protection of vegetation. The 2010 proposal for secondary O3 standard (75 FR 2938,
26 p. 3003) proposed to use the maximum consecutive 3-month period within the O3 season.
27 The U.S. Forest Service and federal land managers have used a 24-h W126 accumulated
28 for 6 months from April through September (2000). However, some monitors in the U.S.
29 are operational for as little as four months and would not have enough data for a 6-month
30 seasonal window. The exposure period in the vast majority of O3 exposure studies
31 conducted in the U.S. has been much shorter than 6 months. Most of the crop studies
32 done through NCLAN had exposures less than three months with an average of 77 days.
33 Open-top chamber studies of tree seedlings, compiled by the EPA, had an average
34 exposure of just over three months or 99 days. In more recent FACE experiments,
35 Soy FACE exposed soybeans for an average of approximately 120 days per year and the
36 Aspen FACE experiment exposed trees to an average of approximately 145 days per year
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1
2
3
of elevated O3, which included the entire growing season at those particular sites. Despite
the possibility that plants may be exposed to ambient O3 longer than 3 months in some
locations, there is a lack of exposure experiments conducted for longer than 3 months.
B
30 40
Highest 3 month W126
30 40 50
Highest 3 month W126
Figure 9-13 Maximum 3-month, 12-h W126 plotted against maximum 6-month,
12-h W126. Data are from the AQS and CASTNET monitors for the
years 2008 and 2009. (A) W126, 3 month versus 6 month, 2008
(Pearson correlation = 0.99); (B) W126, 3 month versus 6 month,
2009 (Pearson correlation = 0.99).
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1 In an analysis of the 3- and 6-month maximum W126 values calculated for over 1,200
2 AQS (Air Quality System) and CASTNET (Clean Air Status and Trend Network) EPA
3 monitoring sites for the years 2008-2009, it was found that these 2 accumulation periods
4 resulted in highly correlated metrics (Figure 9-13). The two accumulation periods were
5 centered on the yearly maximum for each monitoring site, and it is possible that this
6 correlation would be weaker if the two periods were not temporally aligned. In the U.S.,
7 W126 cumulated over 3 months, and W126 cumulated over 6 months are proxies of one
8 another, as long as the period in which daily W126 is accumulated corresponds to the
9 seasonal maximum. Therefore, it is expected that either statistic will predict vegetation
10 response equally well. In other words, the strength of the correlation between maximum
11 3-month W126 and maximum 6-month W126 is such that there is no material difference
12 in their predictive value for vegetation response.
9.5.4 Ozone Uptake/Dose Modeling for Vegetation
13 Another approach for improving risk assessment of vegetation response to ambient O3 is
14 based on estimating the O3 concentration from the atmosphere that enters the leaf (i.e.,
15 flux or deposition). Interest has been increasing in recent years, particularly in Europe, in
16 using mathematically tractable flux models for O3 assessments at the regional, national,
17 and European scale (Matyssek et al.. 2008; Paoletti and Manning. 2007; M and M. 2004;
18 Emberson et al., 2000b; Emberson et al., 2000a). Some researchers have claimed that
19 using flux models can be used to better predict vegetation responses to O3 than exposure-
20 based approaches (Matyssek et al., 2008). However, other research has suggested that
21 flux models do not predict vegetation responses to O3 better than exposure-based models,
22 such as AOT40 (Gonzalez-Fernandez et al., 2010). While some efforts have been made in
23 the U.S. to calculate O3 flux into leaves and canopies (Fares etal.. 2010a: Turnipseed et
24 al.. 2009: Uddling et al.. 2009: Bergweiler et al.. 2008: Hogg et al.. 2007: Grulke et al..
25 2004: GrantzetaL 1997: GrantzetaL 1995), little information has been published
26 relating these fluxes to effects on vegetation. The lack of flux data in the U.S. and the
27 lack of understanding of detoxification processes have made this technique less viable for
28 vulnerability and risk assessments in the U.S.
29 Flux calculations are data intensive and must be carefully implemented. Reducing
30 uncertainties in flux estimates for areas with diverse surface or terrain conditions to
31 within ±50% requires "very careful application of dry deposition models, some model
32 development, and support by experimental observations" (Wesely and Hicks. 2000). As
33 an example, the annual average deposition velocity of O3 among three nearby sites in
34 similar vegetation was found to vary by ±10%, presumably due to terrain (Brook et al.,
35 1997). Moreover, the authors stated that the actual variation was even greater, because
Draft - Do Not Cite or Quote 9-116 September 2011
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1 stomatal uptake was unrealistically assumed to be the same among all sites, and flux is
2 strongly influenced by stomatal conductance (Brook et al.. 1997; Massman and Grantz.
3 1995; Fuentes et al., 1992; Reich. 1987; Leuning et al., 1979). This uptake-based
4 approach to quantify the vegetation impact of O3 requires inclusion of those factors that
5 control the diurnal and seasonal O3 flux to vegetation (e.g., climate patterns, species
6 and/or vegetation-type factors and site-specific factors). The models have to distinguish
7 between stomatal and non-stomatal components of O3 deposition to adequately estimate
8 actual concentration reaching the target tissue of a plant to elicit a response (Uddling et
9 al., 2009). Determining this O3 uptake via canopy and stomatal conductance relies on
10 models to predict flux and ultimately the "effective" flux (Grunhage et al.. 2004;
11 Massman. 2004; Massman et al., 2000). "Effective flux" has been defined as the balance
12 between O3 flux and detoxification processes (Heath et al.. 2009; Musselman and
13 Massman. 1999; Grunhage and Haenel, 1997; Dammgen et al., 1993). The time-
14 integrated "effective flux" is termed "effective dose." The uptake mechanisms and the
15 resistances in this process, including stomatal conductance and biochemical defense
16 mechanisms, are discussed below. The flux-based index is the goal for the "Level II"
17 critical level for assessment of O3 risk to vegetation and ecosystems across Europe
18 (Ashmoreetal..2004a).
19 An important consideration in both O3 exposure and uptake is how the O3 concentration
20 at the top of low vegetation such as, crops and tree seedlings may be lower than the
21 height at which the measurement is taken. Ambient monitor inlets in the U.S. are
22 typically at heights of 3 to 5 meters. During daytime hours, the vertical O3 gradient can
23 be relatively small because turbulent mixing maintains the downward flux of O3. For
24 example, Horvath et al. (1995) calculated a 7% decrease in O3 going from a height of 4
25 meters down to 0.5 meters above the surface during unstable (or turbulent) conditions in
26 a study over low vegetation in Hungary [see section AX3.3.2. of the 2006 O3 AQCD
27 (U.S. EPA. 2006b)1. There have been several studies indicating decreased O3
28 concentrations under tree canopies (Kolb et al.. 1997; Samuelson and Kelly. 1997; Joss
29 andGraber. 1996; Fredericksen et al.. 1995; Lorenzini and Nali. 1995; Enders. 1992;
30 FontanetaL 1992; Neufeld et al.. 1992). In contrast, for forests, measured data may
31 underestimate O3 concentration at the top of the canopy. The difference between
32 measurement height and canopy height is a function of several factors, the intensity of
33 turbulent mixing in the surface layer and other meteorological factors, canopy height and
34 total deposition to the canopy. Some researchers have used deposition models to estimate
35 O3 concentration at canopy-top height based on concentrations at measurement height
36 (Emberson et al.. 2000a). However, deposition models usually require meteorological
37 data inputs that are not always available or well characterized across large spatial scales.
Draft - Do Not Cite or Quote 9-117 September 2011
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1 Soil moisture is a critical factor in controlling O3 uptake through its effect on plant water
2 status and stomatal conductance. In an attempt to relate uptake, soil moisture, and
3 ambient air quality to identify areas of potential risk, available O3 monitoring data for
4 1983 to 1990 were used along with literature-based seedling exposure-response data from
5 regions within the southern Appalachian Mountains that might have experienced O3
6 exposures sufficient to inhibit growth (Lefohn et al.. 1997). In a small number of areas
7 within the region, O3 exposures and soil moisture availability were sufficient to possibly
8 cause growth reductions in some O3 sensitive species (e.g., black cherry). The
9 conclusions were limited, however, because of the uncertainty in interpolating O3
10 exposures in many of the areas and because the hydrologic index used might not reflect
11 actual water stress.
12 The non-stomatal component of plant defenses are the most difficult to quantify, but
13 some studies are available (Heath et al.. 2009; Barnes et al.. 2002; Plochl et al.. 2000;
14 Chen et al.. 1998; Massman and Grantz. 1995). Massman et al. (2000) developed a
15 conceptual model of a dose-based index to determine how plant injury response to O3
16 relates to the traditional exposure-based parameters. The index used time-varying -
17 weighted fluxes to account for the fact that flux was not necessarily correlated with plant
18 injury or damage. The model applied only to plant foliar injury and suggested that
19 application of flux-based models for determining plant damage (yield or biomass) would
20 require a better understanding and quantification of the relationship between injury and
21 damage.
9.5.5 Summary
22 Exposure indices are metrics that quantify exposure as it relates to measured plant
23 damage (i.e., reduced growth). They are summary measures of monitored ambient O3
24 concentrations over time intended to provide a consistent metric for reviewing and
25 comparing exposure-response effects obtained from various studies. No new information
26 is available since 2006 that alters the basic conclusions put forth in the 2006 and 1996 O3
27 AQCDs. These AQCDs focused on the research used to develop various exposure indices
28 to help quantify effects on growth and yield in crops, perennials, and trees (primarily
29 seedlings). The performance of indices was compared through regression analyses of
30 earlier studies designed to support the estimation of predictive O3 exposure-response
31 models for growth and/or yield of crops and tree (seedling) species.
32 Another approach for improving risk assessment of vegetation response to ambient O3 is
33 based on determining the O3 concentration from the atmosphere that enters the leaf (i.e.,
34 flux or deposition). Interest has been increasing in recent years, particularly in Europe, in
Draft - Do Not Cite or Quote 9-118 September 2011
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1 using mathematically tractable flux models for O3 assessments at the regional, national,
2 and European scale (Matyssek et al.. 2008; Paoletti and Manning. 2007; M and M. 2004;
3 Emberson et al., 2000b; Emberson et al., 2000a). While some efforts have been made in
4 the U.S. to calculate O3 flux into leaves and canopies (Turnipseed et al.. 2009; Uddling et
5 al.. 2009; Bergweiler et al.. 2008; Hogg et al.. 2007; Grulke et al.. 2004; Grantz et al..
6 1997; Grantz etal.. 1995). little information has been published relating these fluxes to
7 effects on vegetation. There is also concern that not all O3 stomatal uptake results in a
8 yield reduction, which depends to some degree on the amount of internal detoxification
9 occurring with each particular species. Those species having high amounts of
10 detoxification potential may, in fact, show little relationship between O3 stomatal uptake
11 and plant response (Musselman and Massman. 1999). The lack of data in the U.S. and the
12 lack of understanding of detoxification processes have made this technique less viable for
13 vulnerability and risk assessments in the U.S.
14 The main conclusions from the 1996 and 2006 O3 AQCDs regarding indices based on
15 ambient exposure are still valid. These key conclusions can be restated as follows:
16 • O3 effects in plants are cumulative;
17 • higher O3 concentrations appear to be more important than lower
18 concentrations in eliciting a response;
19 • plant sensitivity to O3 varies with time of day and plant development stage;
20 and
21 • exposure indices that cumulate hourly O3 concentrations and preferentially
22 weight the higher concentrations have better statistical fits to growth/yield
23 response data than do the mean and peak indices.
24 Various weighting functions have been used, including threshold-weighted (e.g.,
25 SUM06) and continuous sigmoid-weighted (e.g., W126) functions. Based on statistical
26 goodness-of-fit tests, these cumulative, concentration-weighted indices could not be
27 differentiated from one another using data from previous exposure studies. Additional
28 statistical forms for O3 exposure indices have been discussed in Lee et al. (1988b). The
29 majority of studies published since the 2006 O3 AQCD do not change earlier
30 conclusions, including the importance of peak concentrations, and the duration and
31 occurrence of O3 exposures in altering plant growth and yield.
32 Given the current state of knowledge and the best available data, exposure indices that
33 cumulate and differentially weight the higher hourly average concentrations and also
34 include the mid-level values continue to offer the most defensible approach for use in
35 developing response functions and comparing studies, as well as for defining future
3 6 indice s for vegetation protection.
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9.6 Ozone Exposure-Plant Response Relationships
9.6.1 Introduction
1 The adequate characterization of the effects of O3 on plants for the purpose of setting air
2 quality standards is contingent not only on the choice of the index used (i.e. SUM06,
3 W126) to summarize O3 concentrations (Section 9.5), but also on quantifying the
4 response of the plant variables of interest at specific values of the selected index. The
5 many factors that determine the response of plants to O3 exposure have been discussed in
6 previous sections. They include species, genotype and other genetic characteristics
7 (Section 9.3), biochemical and physiological status (Section 9.3), previous and current
8 exposure to other stressors (Section 9.4.8), and characteristics of the exposure itself
9 (Section 9.5). Establishing a secondary air quality standard requires the capability to
10 generalize those observations, in order to obtain predictions that are reliable enough
11 under a broad variety of conditions, taking into account these factors. This section
12 reviews results that have related specific quantitative observations of O3 exposure with
13 quantitative observations of plant responses, and the predictions of responses that have
14 been derived from those observations through empirical models.
15 For four decades, exposure to O3 at ambient concentrations found in many areas of the
16 U.S. has been known to cause detrimental effects in plants (U.S. EPA. 2006b. 1996b.
17 1984. 1978a). Results published after the 2006 O3 AQCD continue to support this
18 finding, and the following sections deal with the quantitative characterizations of the
19 relationship, and what new insights may have appeared since 2006. Detrimental effects
20 on plants include visible injury, decreases in the rate of photosynthesis, reduced growth,
21 and reduced yield of marketable plant parts. Most published exposure-response data have
22 reported O3 effects on the yield of crops and the growth of tree seedlings, and those two
23 variables have been the focus of the characterization of ecological impacts of O3 for the
24 purpose of setting secondary air quality standards. In order to support quantitative
25 modeling of exposure-response relationships, data should preferably include more than
26 three levels of exposure, and some control of potential confounding or interacting factors
27 should be present in order to model the relationship with sufficient accuracy. Letting
28 potential confounders, such as other stressors, vary freely when generating O3 exposure-
29 response data might improve the 'realism' of the data, but it also greatly increases the
30 amount of data necessary to extract a clear quantitative description of the relationship.
31 Conversely however, experimental settings should not be so exhaustively restrictive as to
32 make generalization outside of them problematic. During the last four decades, many of
33 the studies of the effects of O3 on growth and yield of plants have not included enough
34 levels of O3 to parameterize more than the simplest linear model. The majority of these
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1 studies have only contrasted two levels, ambient and elevated, or sometimes three by
2 adding carbon filtration in OTC studies, with little or no consideration of quantitatively
3 relating specific values of exposure to specific values of growth or yield. This is not to
4 say that studies that did not include more than two or three levels of O3 exposure, or
5 studies that were conducted in uncontrolled environments, do not provide exposure-
6 response information that is highly relevant to reviewing air quality standards. In fact,
7 they can be essential in verifying the agreement between predictions obtained through the
8 empirical models derived from experiments such as NCLAN, and observations. The
9 consensus of model predictions and observations from a variety of studies conducted in
10 other locations, at other times, and using different exposure methods, greatly increases
11 confidence in the reliability of both. Furthermore, if they are considered in the aggregate,
12 studies with few levels of exposure or high unaccounted variability can provide
13 additional independent estimates of decrements in plant growth and yield, at least within
14 a few broad categories of exposure.
15 Extensive exposure-response information on a wide variety of plant species has been
16 produced by two long-term projects that were designed with the explicit aim of obtaining
17 quantitative characterizations of the response of such an assortment of crop plants and
18 tree seedlings to O3 under North American conditions: the NCLAN project for crops, and
19 the EPA National Health and Environmental Effects Research Laboratory, Western
20 Ecology Division tree seedling project (NHEERL/WED). The NCLAN project was
21 initiated by the EPA in 1980 primarily to improve estimates of yield loss under field
22 conditions and to estimate the magnitude of crop losses caused by O3 throughout the U.S.
23 (HecketaL 1991; Hecketal.. 1982). The cultural conditions used in the NCLAN studies
24 approximated typical agronomic practices, and the primary objectives were: (1) to define
25 relationships between yields of major agricultural crops and O3 exposure as required to
26 provide data necessary for economic assessments and development of O3 NAAQS; (2) to
27 assess the national economic consequences resulting from O3 exposure of major
28 agricultural crops; and (3) to advance understanding of cause-and-effect relationships that
29 determine crop responses to pollutant exposures.
30 NCLAN experiments yielded 54 exposure-response curves for 12 crop species, some of
31 which were represented by multiple cultivars at several of 6 locations throughout the U.S.
32 The NHEERL/WED project was initiated by EPA in 1988 with the same objectives for
33 tree species, and yielded 49 exposure-responses curves for multiple genotypes of 11 tree
34 species grown for up to three years in Oregon, Michigan, and the Great Smoky
35 Mountains National Park. Both projects used OTCs to expose plants to three to five
36 levels of O3. Eight of the 54 crop datasets were from plants grown under a combination
37 of O3 exposure and experimental drought conditions. Figure 9-14 through 9-17
3 8 summarize some of the NCLAN and NHEERL/WED results.
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1 It should be noted that data from FACE experiments might also be used for modeling
2 exposure-response. They only use two levels of O3 (ambient concentration at the site and
3 a multiple of it), but given that the value of both levels of exposure changes every year,
4 and that they are typically run for many consecutive years, aggregating data over time
5 produces twice as many levels of O3 as there are years. As described in Section 9.2.4,
6 FACE experiments seek to impose fewer constraints on the growth environment then
7 OTCs. As a consequence, FACE studies have to contend with larger variability,
8 especially year-to-year variability, but the difference in experimental conditions between
9 the two methodologies makes comparisons between their results especially useful.
10 Growth and yield of at least one crop (soybean) has been investigated in yearly
11 experiments since 2001 at a FACE facility in Illinois (University of Illinois. 2010;
12 Morgan et al.. 2006). however almost all analyses of SoyFACE published so far have
13 been based on subsets of one or two years, and have only contrasted ambient versus
14 elevated O3 as categorical variables. They have not modeled the response of growth and
15 yield to O3 exposure continuously over the range of exposure values that have occurred
16 over time. The only exception is a study by Betzelberger et al. (2010). who used a linear
17 regression model on data pooled over 2 years. Likewise, trees of three species (trembling
18 aspen, paper birch, and sugar maple) were grown between 1998 and 2009 in a FACE
19 experiment located in Rhinelander, Wisconsin (Pregitzer et al.. 2008; Dickson et al..
20 2000). The Aspen FACE experiment has provided extensive data on responses of trees
21 beyond the seedling stage under long-term exposure, and also on ecosystem-level
22 responses (Section 9.4), but the only attempt to use those data in a continuous model of
23 the response of tree growth to O3 exposure (Percy et al.. 2007) suffered severe
24 methodological problems, some of which are discussed in Section 9.6.3. Finally, one
25 experiment was able to exploit a naturally occurring gradient of O3 concentrations to fit a
26 linear regression model to the growth of cottonwood (Gregg et al.. 2006. 2003). Factors
27 such as genotype, soil type and soil moisture were under experimental control, and the
28 authors were able to partition out the effects of potential confounders such as
29 temperature, atmospheric N deposition, and ambient CO2.
30 A serious difficulty in assessing results of exposure-response research is the multiplicity
31 of O3 metrics that have been used in reporting. As described in Section 9.5, metrics that
32 entail either weighting or thresholding of hourly values cannot be converted into one
33 another, or into unweighted metrics such as hourly average. When computing O3
34 exposure using weighted or thresholded metrics, the computation of each metric has to
3 5 start with the original hourly data. Comparisons of exposure-response models can only be
36 made between studies that used the same metric, and the value of exposure at which a
37 given plant response is expected using one metric of exposure cannot be exactly
38 converted to another metric. Determining the exposure value at which an effect would be
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1 observed in a different metric can only be accomplished by first computing the
2 experimental exposures in this metric from the hourly data, then estimating (fitting)
3 model coefficients again. This problem is irremediable, although useful comparisons
4 might be made using categorical exposures such as 'current ambient exposure' or '2050
5 projected exposure', which can serve as a common reference for quantitative values
6 expressed in various metrics. Studies that contained growth or yield exposure-response
7 data at few levels of exposure, and/or using metrics other than W126 are summarized in
8 Tables 9-18 and 9-19.
9.6.2 Estimates of Crop Yield Loss and Tree Seedling Biomass Loss in the
1996 and 2006 Ozone AQCDs
9 The 1996 and 2006 O3 AQCDs relied extensively on analyses of NCLAN and
10 NHEERL/WED by Lee et al. (1994: 1989. 1988b. 1987). Hogsett et al. (1997). Lee and
11 Hogsett (1999). Heck et al. (1984). Rawlings and Cure (1985). Lesser et al. (1990). and
12 Gumpertz and Rawlings (1992). Those analyses concluded that a three-parameter
13 Weibull model-
Equation 9-2
14 is the most appropriate model for the response of absolute yield and growth to O3
15 exposure, because of the interpretability of its parameters, its flexibility (given the small
16 number of parameters), and its tractability for estimation. In addition, removing the
17 intercept a results in a model of relative yield (yield relative to [yield at exposure=0])
18 without any further reparameterization. Formulating the model in terms of relative yield
19 or relative yield loss (yield loss=[l - relative yield]) is essential in comparing exposure-
20 response across species, genotypes, or experiments for which absolute values of the
21 response may vary greatly. In the 1996 and 2006 O3 AQCDs, the two-parameter model
22 of relative yield was used in deriving common models for multiple species, multiple
23 genotypes within species, and multiple locations.
24 Given the disparate species, genotypes, and locations that were included in the NCLAN
25 and NHEERL/WED projects, and in the absence of plausible distributional assumptions
26 with respect to those variables, a three step process using robust methods was used to
27 obtain parameter estimates that could be generalized. The models that were derived for
28 each species or group of species were referred to as median composite functions. In the
29 first step, the three parameters of the Weibull model were computed for absolute yield or
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1 biomass data from each NCLAN and NHEERL/WED experiment (54 crop datasets and
2 49 tree seedling datasets), using nonlinear regression. When data were only available for
3 three levels of exposure because of experimental problems, the shape parameter (3 was
4 constrained to 1, reducing the model to an exponential decay model. In the second step, a
5 was dropped, and predicted values of relative yield or biomass were then computed for
6 12-hr W126 exposures between 0 and 60 ppm-h. At each of these W126 exposure values,
7 the 25th, 50th, and 75th percentiles of the response were identified among the predicted
8 curves of relative response. For example, for the 34 NCLAN studies of 12 crop species
9 grown under non-droughted conditions for a complete cropping cycle (Figure 9-14), the 3
10 quartiles of the response were identified at every integer value of W126 between 0 and
11 60. The third step fitted a two-parameter Weibull model to those percentiles, yielding the
12 median composite function for the relative yield or biomass response to O3 exposure for
13 each grouping of interest (e.g., all crops, all trees, all datasets for one species), as well as
14 composite functions for the other quartiles. In the 1996 and 2006 O3 AQCDs this
15 modeling of crop yield loss and tree seedling biomass loss was conducted using the
16 SUM06 metric for exposure. This section updates those results by using the 12-hr W126
17 as proposed in 2007 (72 FR 37818) and 2010 (75 FR 2938, p. 3003). Figures 9-14
18 through 9-17 present quantiles of predicted relative yield or biomass loss at seven values
19 of the 12-h W126 for some representative groupings of NCLAN and NHEERL/WED
20 results. Tables 9-10 through 9-12 give the 90-day 12-h W126 O3 exposure values at
21 which 10 and 20% yield or biomass losses are predicted in 50 and 75% of crop or tree
22 species using the composite functions.
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100 i
90 •
80 •
g 70 •
w
8 60 •
1 50 •
1 40 •
S. 30 •
20 •
10 •
n .
34 crop datasets
jj
H
•--
4
•
j 90thPctile
T
^*
\
^
?-*
^
\
M i
75thPctile
50thPctile
25thPctile
10thPctile
10 20 30 40 50
12 hrW126 (ppm-hr)
60
Source of Weibull parameters: Lee and Hogsett (1996).
Quantiles of the predicted relative yield loss at 7 values of 12-hour W126 for 34 Weibull curves estimated using nonlinear
regression on data from 34 studies of 12 crop species grown under well-watered conditions for the full duration of 1 cropping cycle.
Figure 9-14 Quantiles of predicted relative yield loss for 34 NCLAN crop
experiments.
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11 Soybean datasets
10 20 30 40 50 60
100 -
90 -
80 -
70 -
60 -
50 -
40 -
30 -
20 -
10 -
5 Cotton datasets
20 30 40 50
12hrW126 (ppm-hr
100 -
90 -
80 -
70 -
60 -
50 -
40 -
30 -
20 -
10 -
0
2 Com datasets
20 30 40 50
12hrW126 (ppm-hr)
Source of Weibull parameters: Lee and Hogsett (1996).
Figure 9-15 Quantiles of predicted relative yield loss for 4 crop species in
NCLAN experiments. Quantiles of the predicted relative yield loss
at 7 values of 12-h W126 for Weibull curves estimated using
nonlinear regression for 4 species grown under well-watered
conditions for the full duration of 1 cropping cycle. The number of
studies available for each species is indicated on each plot.
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9-126
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in
in
o
_i
in
in
ra
o
In
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100
90
80
I ™
I 6°
1 50
o
to 40
s
te 30
Q-
20
10
0
14 Aspendatasets
^- SCPPctile
100 -
90 -
80 -
70 -
60 -
50 -
40 -
30 -
20 -
10 -
0
11 Ponderosapine datasets
10 20 30 40 50
100 -
90 -
80 -
50-
O
™ 40-
I 30 -
Q_
20 -
10 -
0
7 Douglas firdatasets
20 30 40 50
90 day 12 hr W126 (ppm-hr)
100 -
90 -
80 -
70 -
60 -
50 -
40 -
30 -
20 -
10 -
STulip poplardatasets
10 20 30 40 50
90 day 12 hr W126 (ppm-hr)
Source of Weibull parameters: Lee and Hogsett (1996).
Figure 9-17 Quantiles of predicted relative biomass loss for 4 tree species in
NHEERL/WED experiments. Quantiles of the predicted relative
above-ground biomass loss at 7 exposure values of 12-h W126 for
Weibull curves estimated using nonlinear regression on data for 4
tree species grown under well-watered conditions for 1 or 2 year.
Curves were standardized to 90-day W126. The number of studies
available for each species is indicated on each plot.
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Table 9-9 Ozone exposures at which 10 and 20% yield loss is predicted for 50
and 75% of crop species, based on composite functions for the
50th and 75th percentiles of 34 Weibull curves for relative yield loss
data from 34 non-droughted NCLAN studies of 12 crop species;
curves were standardized to 90-day W126
90-day 12-h W126 for 10% yield loss 90-day 12-h W126 for 20% yield loss
(Ppm-h) (ppm-h)
Model for the 50th Percentile of 34 curves
Relative yield=exp(-(W126/104.82)**1.424) 22 37
Model for the 75th Percentile of 34 curves
Relative yield=exp(-(W126/78.12)**1.415) 16 27
Source of parameters for the 34 curves: Lee and Hogsett (1996)
Table 9-10 Ozone exposures at which 10 and 20% yield loss is predicted for 50
and 75% of crop species under drought conditions and adequate
moisture, based on composite functions for the 50th and 75th
percentiles of 16 Weibull curves for relative yield loss data from 8
NCLAN studies that paired droughted and watered conditions for
the same genotype; curves were standardized to 90-day W126
90day12-hW126for10% 90 day 12-h W126 for 20%
yield loss (ppm-h) yield loss (ppm-h)
Model for the 50th Percentile of 2x8 curves
Watered Relative yield=exp(-(W126/132.86)**1.170) 19 37
Droughted Relative yield=exp(-(W126/179.84)**1.713) 48 75
Model for the 75th Percentile of 2x8 curves
Watered Relative yield=exp(-(W126/90.43)**1.310) 16 29
Droughted Relative yield=exp(-(W126/105.16)"1.833) 31 46
Source of parameters for the 16 curves: Lee and Hogsett (1996)
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Table 9-11 Ozone exposures at which 10 and 20% biomass loss is predicted
for 50 and 75 %of tree species, based on composite functions for
the 50th and 75th percentiles of 49 Weibull curves for relative
above-ground biomass loss data from 49 studies of 11 tree species
grown under well-watered conditions for 1 or 2 year; curves were
standardized to 90-day W126
90 day 12 h W126 for 10% yield 90 day 12 h W126 for 20% yield
loss (ppm-h) loss (ppm-h)
Model for the 50th Percentile of 49 curves
Relative yield=exp(-(W126/131.57)**! .242) 21 39
Model for the 75th Percentile of 49 curves
Relative yield=exp(-(W126/65.49)**! .500) 15 24
Source of parameters for the 49 curves: Lee and Hogsett (1996)
9.6.3 Validation of 1996 and 2006 Ozone AQCD Models and Methodology
Using the 90 day 12-h W126 and Current FACE Data
1 Since the completion of the NCLAN and NHEERL/WED projects, almost no studies
2 have been published that could provide a basis for estimates of exposure-response that
3 can be compared to those of the 1996 and 2006 O3 AQCDs. Most experiments,
4 regardless of exposure methodology, include only two levels of exposure. In addition,
5 very few studies have included measurements of exposure using the W126 metric, or the
6 hourly O3 concentration data that would allow computing exposure using the W126. Two
7 FACE projects, however, were conducted over multiple years, and by adding to the
8 number of exposure levels over time, may support independent model estimation and
9 prediction using the same model and the same robust process as summarized in Section
10 9.6.2. Hourly O3 data were available from both FACE projects.
11 The SoyFACE project is situated near Champaign, IL, and comprises 32 octagonal rings
12 (20m-diameter), 4 of which in a given year are exposed to ambient conditions, and 4 of
13 which are exposed to elevated O3 as a fixed proportion of the instantaneous ambient
14 concentration (Betzelberger et al., 2010; University of Illinois. 2010; Morgan et al., 2006;
15 Morgan et al.. 2004). Since 2002, yield data have been collected for up to 8 genotypes of
16 soybean grown in subplots within each ring. The Aspen FACE project is situated in
17 Rhinelander, WI, and comprises 12 rings (30m-diameter), 3 of which are exposed to
18 ambient conditions, and 3 of which are exposed to O3 as a fixed proportion of the
19 instantaneous ambient concentration (Pregitzer et al., 2008; Karnosky et al., 2005;
20 Dickson et al.. 2000). In the summer of 1997, half the area of each ring was planted with
21 small (five to seven leaf sized) clonally propagated plants of five genotypes of trembling
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1 aspen, which were left to grow in those environments until 2009. Biomass data are
2 currently available for the years 1997-2005 (King et al.. 2005). Ozone exposure in these
3 two FACE projects can be viewed as a categorical variable with two levels: ambient, and
4 elevated. However, this overlooks the facts that yearly ambient and elevated exposure
5 both vary with every year, and that the proportionality between them also changes. This
6 change has two sources: first, the dispensing of O3 into the elevated exposure rings varies
7 from the proportionality set point to some extent, and for SoyFACE, the set point
8 changed between years. Second, the proportionality does not propagate predictably from
9 the hourly data to the yearly value when using threshold or concentration-weighted
10 cumulative metrics (such as AOT40, SUM06 or W126). Hourly average elevated
11 exposures that are, for example, a constant 1.5 times greater than ambient do not result in
12 AOT40, SUM06 or W126 values that are some constant multiple of the ambient values of
13 those indices. The greater the fraction of elevated hourly values that are above the
14 threshold or heavily weighted, compared to the fraction of hourly ambient values that are,
15 the greater the difference between ambient and elevated yearly exposure, as measured
16 using weighted cumulative indices. When elevated exposure is a multiple of ambient
17 hourly intervals, the number of hours for which elevated exposure meets the threshold for
18 inclusion can vary widely, even though the hourly mean for the year retains the
19 proportionality. As a consequence, the number of exposure levels in multi-year
20 experiments is twice the number of years. In the case of SoyFACE for the period between
21 2002 and 2008, ambient exposure in the highest year was approximately equal to elevated
22 exposure in the lowest year, with 14 levels of O3 exposure evenly distributed from lowest
23 to highest. The particular conditions of the Aspen FACE experiment resulted in 12
24 exposure levels between 1998 and 2003, but they were not as evenly distributed between
25 minimum and maximum over the 6-year period.
26 There are necessary differences in the modeling of exposure-response in annual plants
27 such as soybean, and in perennial plants such as aspen trees, when exposure takes place
28 over multiple years. In annual plants, responses recorded at the end of the life cycle, i.e.,
29 yearly, are analyzed in relationship to that year's exposure. Yield of soybeans is affected
30 by exposure during the year the crop was growing, and a new crop is planted every year.
31 Thus an exposure-response relationship can be modeled from yearly responses matched
32 to yearly exposures, with those exposure-response data points having been generated in
33 separate years. For perennial organisms, which are not harvested yearly and continue to
34 grow from year to year, such pairing of exposure and response cannot be done without
3 5 accounting for time. Not only does the size of the organism at the beginning of each year
36 of exposure increase, but size is also dependent on the exposure from previous years.
37 Therefore the relationship of response and exposure must be analyzed either one year at a
38 time, or by standardizing the response as a yearly increment relative to size at the
39 beginning of each year. Furthermore, the relevant measurement of exposure is
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1 cumulative, or cumulative yearly average exposure, starting in the year exposure was
2 initiated, up to the end of the year of interest. When analyzing the growth of trees over
3 several years, it would be evidently incorrect to pair the exposure level in every discrete
4 year with absolute size of the trees that year, and posit a direct relationship between them,
5 without taking increasing age into consideration. In the Aspen FACE experiment, for
6 example, one could not establish an exposure-response relationship by matching
7 12 yearly exposures and 12 yearly tree sizes, while disregarding age as if size did not also
8 depend on it. This is the basis of the 2007 study of Aspen FACE data by Percy et al.
9 (2007). which compares the size of trees of various ages as if they were all the same age,
10 and was therefore not informative.
9.6.3.1 Comparison of NCLAN-Based Prediction and SoyFACE
Data
11 For this ISA, EPA conducted a comparison between yield of soybean as predicted by the
12 composite function three-step process (Section 9.6.2) using NCLAN data, and
13 observations of yield in SoyFACE. The median composite function for relative yield was
14 derived for the 11 NCLAN soybean Weibull functions for non-droughted studies, and
15 comparisons between the predictions of the median composite and SoyFACE
16 observations were conducted as follows.
17 For the years 2007 and 2008, SoyFACE yield data were available for 7 and 6 genotypes,
18 respectively. The EPA used those data to compare the relative change in yield observed
19 in SoyFACE in a given year between ambient O3 and elevated O3, versus the relative
20 change in yield predicted by the NCLAN-based median composite function between
21 those same two values of O3 exposure. The two parameter median composite function for
22 relative yield of soybean based on NCLAN data was used to predict yield response at the
23 two values of exposure observed in SoyFACE in each year, and the change between yield
24 under ambient and elevated was compared to the change observed in SoyFACE for the
25 relevant year (Table 9-12). This approach results in a direct comparison of predicted
26 versus observed change in yield. Because the value of relative response between any two
27 values of O3 exposure is independent of the intercept a, this comparison does not require
28 prediction of the absolute values of the responses.
29 Since comparisons of absolute values might be of interest, the predictive functions were
30 also scaled to the observed data: SoyFACE data were used to compute an intercept a
31 while the shape and scale parameters (|3 and r\) were held at their value in the NCLAN
32 predictive model. This method gives a comparison of prediction and observation that
33 takes all the observed information into account to provide the best possible estimate of
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1
2
3
4
5
6
7
9
10
11
the intercept, and thus the best possible scaling (Table 9-13 and Figure 9-18). For the
comparison of NCLAN and SoyFACE, this validation was possible for 2007 and 2008,
where data for 7 and 6 soybean genotypes, respectively, were available. The median
composite function for relative yield was derived for the 11 NCLAN soybean Weibull
functions for nondroughted studies, and the values of median yield under ambient
exposure at SoyFACE in 2007 and 2008 were used to obtain an estimate of the intercept
a for the NCLAN median function in each of the two years.
Table 9-12 presents the results of ambient/elevated relative yield comparisons between
the NCLAN-derived predictions and SoyFACE observations. Table 9-13 and figure 9-18
present the results of comparisons between NCLAN-derived predictions and SoyFACE
observations of yield, with the predictive function scaled to provide absolute yield values.
Table 9-12 Comparison between change in yield observed in the SoyFACE
experiment between elevated and ambient ozone, and change
predicted at the same values of ozone by the median composite
function for NCLAN (two-parameter relative yield model)
Year
90-day 12-hW1 26 (ppm-h)
observed at SoyFACE
Ambient Elevated
2007
2008
4.39 46.23
3.23 28.79
Yield in
Elevated O3 Relative to Ambient O3 (%)
Predicted by NCLAN Observed at SoyFACE
75 76
85 88
Table 9-13 Comparison between yield observed in the SoyFACE experiment
and yield predicted at the same values of ozone by the median
composite function for NCLAN (three-parameter absolute yield
model with intercept scaled to SoyFACE data)
Year
90-day 12-h W126 (ppm-h)
observed at SoyFACE
Yield predicted by NCLAN (g/m2)
Yield observed at SoyFACE
(g/m2)
Ambient
Elevated
Ambient
Elevated
Ambient
Elevated
2007
4.39
46.23
309.2
230.6
305.2
230.6
2008
3.23
28.79
350.3
298.2
344.8
304.4
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400 -
350 -
300 -
250 -
200 -
150 -
100 -
50 -
0 -
2007, 7 genotypes
400
350
300
_ 250
^E
3 200 -
1
*~ 150 -
100 -
50 -
0
2008, 6 genotypes
10 20 30 40 50
90day12hrW126 (ppm-hr)
20 30 40 50
90 day 12 hr W126 (ppm-hr)
Source of data: Betzelberger et al. (2010): Morgan et al. (2006): Lee and Hogsett (1996).
Note: Black dots are the median of 7 or 6 soybean genotypes in SoyFACE (2007, 2008); bars are IQR for genotypes; dashed line
is median composite model for 11 studies in NCLAN.
Figure 9-18 Comparison of yield observed in SoyFACE experiment in a given
year with yield predicted by the median composite function based
on NCLAN.
1
2
3
4
5
6
7
8
9
10
11
12
13
Finally, a composite function for the 25th, 50th, and 75th percentiles was developed from
SoyFACE annual yield data, and compared to the NCLAN-based function. The process
described in Section 9.6.2 was applied to SoyFACE data for individual genotypes,
aggregated over the years during which each was grown; one genotype from 2003 to
2007, and six genotypes in 2007 and 2008. First, the three parameter Weibull model
described in Section 9.6.2 was estimated using nonlinear regression on exposure-yield
data for each genotype separately, over the years for which data were available, totaling
seven curves. The 25th, 50th, and 75th percentiles of the predicted values for the two
parameter relative yield curves were then identified at every integer of W126 between 0
and 60, and a two-parameter Weibull model estimated by regression for the three
quartiles. The comparison between these composite functions for the quartiles of relative
yield loss in SoyFACE and the corresponding composite functions for NCLAN is
presented in Figure 9-19.
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100
90 •
80 •
g 70 •
(/)
§ 60 •
1 50 •
| 40 -
I 30
20
10
0 -I
FACE75thPctile
FACE 25th Pctile
FACE median
10 20 30 40 50
90 day 12hrW126 (ppm-hr)
60
70
Source of data: Betzelberger et al. (2010); Morgan et al. (2006): Lee and Hogsett (1996).
Figure 9-19 Comparison of composite functions for the quartiles of 7 curves for
7 genotypes of soybean grown in the SoyFACE experiment, and for
the quartiles of 11 curves for 5 genotypes of soybean grown in the
NCLAN project.
1
2
3
4
5
6
1
8
9
10
11
12
13
14
15
16
As seen in Tables 9-13 and 9-14, and in Figure 9-18, the agreement between predictions
based on NCLAN data and SoyFACE observations was notably close in single-year
comparisons. Together with the very high agreement between median composite models
for NCLAN and SoyFACE (Figure 9-19), it provides very strong mutual confirmation of
those two projects' results with respect to the response of yield of soybeans to O3
exposure. It is readily apparent from these results that the methodology described in
Section 9.6.2 for obtaining predictions of yield or yield loss from NCLAN data is
strongly validated by SoyFACE results. As described in Section 9.2, the exposure
technologies used in the two projects were in sharp contrast, specifically with respect to
the balance each achieved between control of potential interacting factors or confounders,
and fidelity to natural conditions. The comparisons that EPA conducted therefore
demonstrate that the methodology used in developing the composite functions is resistant
to the influence of nuisance variables and that predictions are reliable. They may also
suggest that the aspects in which the two exposure technologies differ have less influence
on exposure-response than initially supposed. These results are also in agreement with
comparative studies reviewed in 9.2.6.
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9.6.3.2 Comparison of NHEERL/WED-Based Prediction of Tree
Biomass Response and Aspen FACE Data
1 EPA also conducted two comparisons between prediction of above-ground biomass loss
2 based on NHEERL/WED results and observations from Aspen FACE. The median
3 composite function was developed from NHEERL/WED data for 11 studies that used
4 wild-type seedlings of aspen as well as four clonally propagated genotypes. All plants
5 were grown in OTCs for one growing season before being destructively harvested. Aspen
6 FACE data were from clonally propagated trees of five genotypes grown from 1998 to
7 2003, with above-ground biomass calculated using allometric equations derived from
8 data for trees harvested destructively in 2000 and 2002 (King et al.. 2005).
9 The two parameter median composite function for relative biomass was used to predict
10 biomass response under the observed elevated exposure, relative to its value under
11 observed ambient exposure, for each separate year of Aspen FACE. EPA first compared
12 Aspen FACE observations of the change in biomass between ambient and elevated
13 exposure with the corresponding prediction at the same values of exposure. Comparisons
14 between observed and predicted absolute biomass values were then conducted for each
15 year by scaling the predictive function to yearly Aspen FACE data as described for
16 soybean data in Section 9.6.3.1. In all cases, yearly 90 day 12-hour W126 values for
17 Aspen FACE were computed as the cumulative average from the year of planting up to
18 the year of interest. A comparison of composite functions between NHEERL/WED and
19 Aspen FACE, similar to the one performed for NCLAN and SoyFACE, was not possible:
20 as discussed in the introduction to Section 9.6, the pairing of 12 exposure values from
21 separate years and 12 values of biomass cannot be the basis for a model of exposure-
22 response, because the trees continued growing for the six-year period of exposure.
23 Because the same trees were used for the entire duration, and continued to grow, data
24 could not be aggregated over years. Table 9-14 presents the results of ambient/elevated
25 relative biomass comparisons between the NHEERL/WED-derived predictions and
26 Aspen FACE observations. Table 9-15 and Figure 9-20 present the results of the
27 comparison between NHEERL/WED-derived predictions and Aspen FACE observations
28 for absolute biomass, using Aspen FACE data to scale the NHEERL/WED-derived
29 composite function.
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Table 9-14 Comparison between change in above-ground biomass elevated
and ambient ozone in Aspen FACE experiment in 6 year, and
change predicted at the same values of ozone by the median
composite function for NHEERL/WED (two-parameter relative
biomass model)
Year
90-day 12-h W126 (ppm-h)
Cumulative Average observed at Aspen FACE
Above-Ground Biomass in
Elevated O$ relative To Ambient O$ (%)
1998
1999
2000
2001
2002
2003
Ambient
3.19
2.61
2.43
2.55
2.51
2.86
Elevated
30.08
33.85
30.16
31.00
30.27
29.12
Predicted by NHEERL/WED
74
70
74
73
74
75
Observed at Aspen FACE
75
70
71
71
69
71
Table 9-15 Comparison between above-ground biomass observed in Aspen
FACE experiment in 6 year and biomass predicted by the median
composite function based on NHEERL/WED (three-parameter
absolute biomass model with intercept scaled to Aspen FACE data)
90day12-hW126(ppm-h)
Year Cumulative Average observed
at Aspen FACE
Biomass Predicted by
NHEERL/WED (g/m"
Biomass Observed, at Aspen
FACE (g/n?)
Ambient
Elevated
Ambient
Elevated
Ambient
Elevated
1998
3.19
30.08
276.0
203.2
274.7
204.9
1999
2.61
33.85
958.7
668.3
955.3
673.3
2000
2.43
30.16
1382.4
1022.8
1400.3
998.6
2001
2.55
31.00
1607.0
1173.7
1620.7
1154.9
2002
2.51
30.27
2079.0
1532.1
2125.9
1468.41
2003
2.86
29.12
2640.1
1981.2
2695.2
1907.8
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3000 n
2500 -
2000 -
•*-...,
1000 -
500 -
-*--. *^
1500 - - - .. „
* ^ •» ^ ^ ••» ^
^ •* ^ ^ ^
" f
"it
2003
2002
2001
2000
1999
1998
10 20 30 40 50 60
90 day 12 hr W126 (yearly cumulative average, ppm-hr)
70
Source of data: King et al. (2005). Lee and Hogsett (1996).
Note: Black dots are aspen biomass/m2 for 3 FACE rings filled with an assemblage of 5 clonal genotypes of aspen at Aspen
FACE; bars are SE for 3 rings; dashed line is median composite model for 4 clonal genotypes and wild-type seedlings in 11
NHEERL/WED 1-year OTC studies.
Figure 9-20 Comparison between above-ground biomass observed in Aspen
FACE experiment in 6 year and biomass predicted by the median
composite function based on NHEERL/WED.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
As in the comparisons between NCLAN and SoyFACE, the agreement between
predictions based on NHEERL/WED data and Aspen FACE observations was very close.
The results of the two projects strongly reinforce each other with respect to the response
of aspen biomass to O3 exposure. The methodology used for obtaining the median
composite function is shown to be capable of deriving a predictive model despite
potential confounders, and despite the added measurement error that is expected from
calculating biomass using allometric equations. In addition, the function based on
one year of growth was shown to be applicable to subsequent years.
The results of experiments that used different exposure methodologies, different
genotypes, locations, and durations converged to the same values of response to O3
exposure for each of two very dissimilar plant species, and predictions based on the
earlier experiments were validated by the data from current ones. However, in these
comparisons, the process used in establishing predictive functions involved aggregating
data over variables such as time, locations, and genotypes, and the use of a robust statistic
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1 (quartiles) for that aggregation. The validating data, from SoyFACE and Aspen FACE,
2 were in turn aggregated over the same variables. The accuracy of predictions is not
3 expected to be conserved for individual values of those variables over which aggregation
4 occurred. For example, the predicted values for soybean, based on data for five
5 genotypes, are not expected to be valid for each genotype separately. As shown in the
6 validation, however, aggregation that occurred over different values of the same variable
7 did not affect accuracy: composite functions based on one set of genotypes were
8 predictive for another set, as long as medians were used for both sets. A study of
9 cottonwood (Populus deltoides) conducted using a naturally occurring gradient of O3
10 exposure (Gregg et al.. 2006. 2003) may provide an illustration of the response of an
11 individual species whose response is far from the median response for an aggregation of
12 species.
9.6.3.3 Exposure-Response in a Gradient Study
13 Gregg et al. (2003) grew saplings of one clonally propagated genotype of cottonwood
14 (Populus deltoides) in seven locations within New York City and in the surrounding
15 region between July and September in 1992, 1993 and 1994, and harvested them 72 days
16 after planting. Owing to regional gradients of atmospheric O3 concentration, the
17 experiment yielded eight levels of exposure (Figure 9-21), and the authors were able to
18 rule out environmental variables other than O3 to account for the large differences in
19 biomass observed after one season of growth. The deficit in growth increased
20 substantially faster with increasing O3 exposure than has been observed in aspen, another
21 species of the same genus (Populus tremuloides, Section 9.6.3.2). Using a three
22 parameter Weibull model (Figure 9-21), the biomass of cottonwood at a W126 exposure
23 of 15 ppm-h, relative to biomass at 5 ppm-h, is estimated to be 0.18 (18% of growth at
24 5 ppm-h). The relative biomass of trembling aspen within the same 5-15 ppm-h range of
25 exposure is estimated to be 0.92, using the median composite model for aspen whose
26 very close agreement with Aspen FACE data was shown in Section 9.6.3.2. Using a
27 median composite function for all deciduous trees in the NHEERL/WED project (6
28 species in 21 studies) also gives predictions that are very distant from the cottonwood
29 response observed in this experiment. For all deciduous tree species in NHEERL/WED,
30 biomass at a W126 exposure of 15 ppm-h, relative to biomass at 5 ppm-h, was estimated
31 to be 0.87.
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10 20 30 40 50
72 day 12 hr W126 (ppm-hr)
60
70
Source: Modified with permission from Nature Publishing Group (Gregg et al.. 2003).
Figure 9-21 Above-ground biomass for one genotype of cottonwood grown in
seven locations for one season in 3 years. Line represents the
three-parameter Weibull model.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
These cottonwood data confirm that, as should be expected, some individual tree species
are substantially more sensitive than the median of NHEERL/WED (Figure 9-16). As
shown in Section 9.6.2, the median models available for trembling aspen and soybean
have verifiable predictive ability for those particular species. This suggests that the
corresponding NCLAN- and NHEERL/WED-based models for multiple crop and tree
species can provide reliable estimates of losses for similar assortments of species.
However, their predictive ability would likely be poor for individual species not tested.
An alternative hypothesis for the difference between the response of cottonwood in this
experiment and deciduous tree species in NHEERL/WED, or the difference between the
response of cottonwood and aspen in NHEERL/WED and Aspen FACE, could be the
presence of confounding factors in the environments where the experiment was
conducted. However, variability in temperature, moisture, soil fertility, and atmospheric
deposition of N were all ruled out by Gregg et al. (2003) as contributing to the observed
response to O3. In addition, this hypothesis would imply that the unrecognized
confounder(s) were either absent from both OTC and FACE studies, or had the same
value in both. This is not impossible, but the hypothesis that cottonwood is very sensitive
to O3 exposure is more parsimonious, and sufficient.
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9.6.3.4 Meta-analyses of growth and yield studies
1
2
3
4
5
6
7
8
9
10
11
12
Since the 2006 O3 AQCD, five studies have used meta-analytic methods to integrate
results from experimental studies of crops or tree species relevant to the U.S. It is
possible to obtain exposure-response data for growth and yield from those meta-analyses,
but because all of them provided summary measurements of O3 exposure as hourly
averages of various lengths of exposures, comparisons with exposure-response results
where exposure is expressed as W126 are problematic. Table 9-16 summarizes the
characteristics of the five meta-analyses. They all included studies conducted in the U.S.
and other locations worldwide, and all of them expressed responses as comparative
change between levels of exposure to O3, with carbon filtered air (CF) among those
levels. Using hourly average concentration to summarize exposure, CF rarely equates
absence of O3, although it almost always near zero when exposure is summarized as
W126, SUM06, or AOT40.
Table 9-1 6
Study
Ainsworth (2008)
Fena et al. (2008b)
Feng and
Kobavashi (2009)
Grantz et al. (2006)
Wittig et al. (2009)
Meta-analyses of growth or yield studies published since
Number of articles
included
12
53
All crops together : 81
16
All responses:263
Articles that included
biomass:unreported
Years pf
1980-2007
1980-2007
1980-2007
1992-2004
1970-2006
Crop, species or
genera
rice
wheat
Potato, barley, wheat, rice,
bean, soybean
34 herbaceous dicots
21 herbaceous monocots
5 tree species
4 gymnosperm tree genera
11 angiosperm tree genera
Response
Yield
Yield
Yield
Relative
Growth Rate
Total
biomass
Number
OfC-3
levels
2
5
3
2
4
2005
Duration of
exposure
unreported
> 10 days
> 10 days
2-24 weeks
> 7 days
13
14
15
16
17
18
19
20
21
22
23
24
The only effect of O3 exposure on yield of rice reported in Ainsworth (2008) was a
decrease of 14% with exposure increasing from CF to 62 ppb average concentration.
Feng et al. (2008b) were able to separate exposure of wheat into four classes with average
concentrations of 42, 69, 97, and 153 ppb, in data where O3 was the only treatment. Mean
responses relative to CF were yield decreases of 17, 25, 49, and 61% respectively. Feng
et al. (2008b) observed that wheat yield losses were smaller under conditions of drought,
and that Spring wheat and Winter wheat appeared similarly affected. However, mean
exposure in studies of Winter wheat was substantially higher than in studies of Spring
wheat (86 versus 64 ppb), which suggests that the yield of Spring wheat was in fact more
severely affected, since yield was approximately the same, even though Spring wheat was
exposed to lower concentrations. Exposures of the six crops considered in Feng and
Kobayashi (2009) were classified into two ranges, each compared to CF air. In the lower
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1 range of exposure (41-49 ppb), potato studies had the highest average exposure (45 ppb)
2 and wheat and rice the lowest (41 ppb). In the higher range (51-75 ppb), wheat studies
3 had the highest average exposure (65 ppb), and potato, barley and rice the lowest (63
4 ppb). In other words, across the studies included, all crops were exposed to very similar
5 levels of O3. At approximately 42 ppb, the yield of potato, barley, wheat, rice, bean, and
6 soybean declined by 5.3, 8.9, 9.7, 17.5, 19, and 7.7% respectively, relative to CF air. At
7 approximately 64 ppb O3, declines were 11.9, 12.5, 21.1, 37.5, 41.4, and 21.6%. Grantz
8 et al. (2006) reported Relative Growth Rate (RGR) rather than growth, and did not report
9 O3 exposure values in a way that would allow calculation of mean exposure for each of
10 the three categories of plants for which RGR changes are reported. All studies used only
11 two levels of exposure, with CF air as the lower one, and most used elevated exposure in
12 the range of 40 to 70 ppb. Decline in RGR was 8.2% for the 34 herbaceous dicots, 4.5%
13 for the 21 herbaceous monocots, and 17.9% for the 5 tree species. Finally, Wittig et al.
14 (2009) divided the studies analyzed into three classes of comparisons: CF versus ambient,
15 CF versus elevated, and ambient versus elevated, but reported comparisons between three
16 average levels of exposure besides CF: 40 ppb, 64 ppb, and 97 ppb. Corresponding
17 decreases in total biomass relative to CF were 7, 17, and 17%.
18 These meta-analyses provide very strong confirmation of EPA's conclusions from
19 previous O3 AQCDs: compared to lower levels of ambient O3, current levels in many
20 locations are having a substantial detrimental effect on the growth and yield of a wide
21 variety of crops and natural vegetation. They also confirm strongly that decreases in
22 growth and yield continue at exposure levels higher than current ambient levels.
23 However, direct comparisons with the predictions of exposure-response models that use
24 concentration-weighted cumulative metrics are difficult.
9.6.3.5 Additional exposure-response data
25 The studies summarized in Tables 9-18 and 9-19 contain growth or yield exposure-
26 response data at too few levels of exposure for exposure-response models, and/or used
27 metrics other than W126. These tables update Tables AX9-16 through AX9-19 of the
28 2006 O3 AQCD.
9.6.4 Summary
29 None of the information on effects of O3 on vegetation published since the 2006 O3
30 AQCD has modified the assessment of quantitative exposure-response relationships that
31 was presented in that document. This assessment updates the 2006 exposure-response
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1 models by computing them using the W126 metric, cumulated over 90 days. Almost all
2 of the experimental research on the effects of O3 on growth or yield of plants published
3 since 2006 used only two levels of exposure. In addition, hourly O3 concentration data
4 that would allow calculations of exposure using the W126 metric are generally
5 unavailable. However, two long-term experiments, one with a crop species (soybean),
6 one with a tree species (aspen), have produced data that can be used to validate the
7 exposure-response models presented in the 2006 O3 AQCD, and methodology used to
8 derive them.
9 Quantitative characterization of exposure-response in the 2006 O3 AQCD was based on
10 experimental data generated for that purpose by the National Crop Loss Assessment
11 Network (NCLAN) and EPA National Health and Environmental Effects Research
12 Laboratory, Western Ecology Division (NHEERL-WED) projects, using OTCs to expose
13 crops and trees seedling to O3. In recent years, yield and growth results for two of the
14 species that had provided extensive exposure-response information in those projects have
15 become available from studies that used FACE technology, which is intended to provide
16 conditions much closer to natural environments (Pregitzer et al.. 2008; Morgan et al..
17 2006; Morgan et al.. 2004; Dickson et al.. 2000). The robust methods that were used
18 previously with exposure measured as SUM06 were applied to the NCLAN and
19 NHEERL-WED data with exposure measured as W126, in order to derive single-species
20 median models for soybean and aspen from studies involving different genotypes, years,
21 and locations. The resulting models were used to predict the change in yield of soybean
22 and biomass of aspen between the two levels of exposure reported in current FACE
23 experiments. Results from these new experiments were exceptionally close to predictions
24 from the models. The accuracy of model predictions for two widely different plant
25 species provides support for the validity of the corresponding multiple-species models for
26 crops and trees in the NCLAN and NHEERL-WED projects. However, variability among
27 species in those projects indicates that the range of sensitivity is likely quite wide. This
28 was confirmed by a recent experiment with cottonwood in a naturally occurring gradient
29 of exposure (Gregg et al., 2006), which established the occurrence of species with
30 responses substantially more severe under currently existing conditions than are predicted
31 by the median model for multiple species.
32 Results from several meta-analyses have provided approximate values for responses of
33 yield of soybean, wheat, rice and other crops under broad categories of exposure, relative
34 to charcoal-filtered air (Ainsworth. 2008; Feng et al.. 2008b; Morgan et al.. 2003).
35 Likewise, Feng and Kobayashi (2009) have summarized yield data for six crop species
36 under various broad comparative exposure categories, while Wittig et al. (2009) reviewed
37 263 studies that reported effects on tree biomass. However, these analyses have proved
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1
2
difficult to compare with exposure-response models, especially given that exposure was
not expressed in the same W126 metric.
Table 9-1 7
Species
Facility
Location
Alfalfa (Medicago
sativa)
OTC; 0.27m3 pots
Federico, Italy
Bean (Phaseolus
vulgaris I . cv
Borlotto)
OTC; ground-
planted
Curno, Italy
Big Blue Stem
(Andropogon
gerardii)
OTC
Alabama
Brassica napus cv.
Westar
Growth chambers
Finland
Corn (Zea mays cv.
Chambord)
OTC
France
Cotton cv. Pima
OTC; 9-L pots
France
Eastern Gamagrass
(Tripsacum
dactybides)
OTC
Alabama
Grapevine (Vitis
vinivera)
OTC
Austria
Mustard (Brassica
campestris)
Chambers;
7.5-cm pots
Oilseed Rape
(Brassica napus)
OTC
Yangtze Delta,
China
Peanut (Arachis
hypogaea)
OTC
Raleigh, NC
Poa pratensis
OTC
Braunschweig,
Germany
Summary of studies of effects of ozone exposure on growth and
yield of agricultural crops
Exposure
Duration
2 yr, 2005,
2006
3 months,
2006
4 months,
2003
17-26 days
33 days
8wk
4 months,
2003
3 yr, May-Oct
10 days
39 days
Syr
2000-2002:
4-5 wk in the
Spring
O3 Exposure
(Additional Treatment)
AOT40: CF 0 ppm-h
1 3.9 ppm-h (2005), 10.1 ppm-h
(2006)
(NaCI: 0.29, 0.65, 0.83,
1 .06 deciSiemens/meter)
Seasonal AOT40:
CF (0.5 ppm-h);
ambient (4.6 ppm-h)
(N/A)
12-havg:
CF(14ppb),
Ambient (29 ppb),
Elevated (71 ppb)
(N/A)
8-h avg:
CF(Oppb), 100 ppb
(Bt/non-Bt; herbivory)
AOT40 ppm-h: 1.1; 1.3; 4.9;
7.2; 9.3; 12.8
(N/A)
1 2-h avg: 12.8 ±0.6; 79.9 ±
6.3; 122.7 ±9.7
(N/A)
12-havg:
CF (14ppb),
Ambient (29 ppb),
Elevated (71 ppb)
(N/A)
AOT40 ppm-h:
CF (0),
Ambient (7-20),
Elevated. 1 (20-30), Elevated.
2 (38-48)
CF&
67.8 ppb for 7 h
(N/A)
Daily avg: 100 ppb, one with
diurnal variation and one with
constant concentration
(N/A)
12-havg:
CF (22 ppb),
Ambient (46 ppb),
Elevated (75ppb)
(C02:375ppm;548ppm;
730 ppm)
8-h avg:
CF+25(21.7),
NF+50(73.1)
(Competition)
Response
Measured
Total shoot yield
# Seeds per plant;
100-seed weight
Final harvest
biomass;
RVF
Shoot biomass
Total above-ground
biomass
Above-ground
biomass
Final harvest
biomass;
RVF
Total fruit yield/
Sugar yield
Seeds/plant
Biomass and pods
per plant
Yield (seed weight,
g/m)
Total biomass (g
DW/pot)
percent change
from CF
(percent change
from ambient)
n.s. (N/A)
-33 (N/A)
n.s. (N/A)
n.s. (n.s.)
-7 (-7)
-30.70 (N/A)
N/A (Highest
treatment caused -
26% change)
-76 (n.s.)
+68 (+42);
-17 (-12)
-20 to -80 in different
yr
(-20 to -90 in different
yr)
n.s. (N/A)
Diurnal variability
reduced both
biomass and pod
number more than
constant fumigation
(N/A)
-33 (-8)
N/A (n.s.)
Reference
Maggio et al.
(2009)
Gerosa et al.
(2009)
Lewis et al.
(2006)
Himanen et al.
(2009b)
Leitao etal.
(2007c)
Grantz and
Shrestha
(2006)
Lewis et al.
(2006)
Sola et al.
(2004)
Black et al.
(2007)
Wang et al.
(2008)
Burkey et al.
(2007)
Bender etal.
(2006)
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Species
Facility
Location
Potato (Solarium
tuberosum)
OTC; CHIP
6 northern European
locations
Rice (Oryza saf/Va)
OTC
Raleigh, NC
Rice (Oryza saf/Va)
20 Asian cultivars
OTC
Gunma Prefecture,
Japan
Seminatural grass
FACE
Le Mouret,
Switzerland
Soybean
OTC; CRA
Bari, Italy
Soybean (Glycine
maxcv. 93B15)
SoyFACE
Urbana, IL
Soybean (Glycine
max cv. Essex)
Chambers; 21 L
Raleigh, NC
Soybean (Glycine
max cv. Essex)
OTCs;21-Lpots
Raleigh, NC
Soybean (Glycine
max cv. Tracaja)
Chambers; pots
Brazil
Soybean (Glycine
max) 10 cultivars
SoyFACE
Urbana, IL
Spring Wheat
(Triticum aestivum
cv. Minaret; Satu;
Drabant; Dragon)
OTCs
Belgium, Finland, &
Sweden
Strawberry (Fragaria
xananassa Duch.
Cv Korona &
Elsanta)
Growth chambers
Bonn, Germany
Sugarbeet (Befa
vulgaris cv. Patriot)
OTC
Belgium
Sugarcane
(Saccharum spp)
CSTR
San Joaquin Valley,
CA
Exposure
Duration
1988,1999.
Emergence to
harvest
1997-1998,
June-
September
2008 growing
season
Syr
2003-2005
growing
seasons
2002, 2003
growing
seasons
2x3 months
2x3 months
20 days
2007 & 2008
1990-2006
2 months
2003, 2004;
5 months
2007;
11-13wk.
Os Exposure
(Additional Treatment)
AOT40:CF (0);
Ambient (0.27-5.19); NF
(0.002-2.93)
NF+ (3.10-24.78
(N/A)
12-hmean ppb:
CF (27.5),
Elevated (74.8)
(CO,
Daily avg (ppb):
CF (2),
O.Sxambient (23);
1 xambient (28);
1.5xambient(42);
2xambient (57)
(Cultivar comparisons)
Seasonal AOT40: Ambient
(0.1-7.2ppm-h);
Elevated. (1.8-24.1 ppm-h)
(N/A)
Seasonal AOT40 ppm-h: CF
(0),
Ambient (3.4), High (9.0)
(Drought)
8-h avg:
Ambient (62 & 50 ppb),
Elevated (75 & 63 ppb)
(N/A)
12-havg:CF(28),
Elevated (79),
Elevated flux (11 2)
(C02: 365 & 700)
12-havg:CF(18);
Elevated (72)
(C02:367&718)
12-havg:CF&30ppb
(N/A)
8-h avg: Ambient (46.3 & 37.9),
Elevated (82.5 & 61 .3)
(Cultivar comparisons)
Seasonal AOT40s ranged from
0 to 16 ppm-h
(N/A)
8-h avg: CF (0 ppb) &
Elevated (78 ppb)
(N/A)
8-h avg: Ambient (36 ppb);
Elevated (62 ppb)
(N/A)
12-havg:CF(4ppb);
Ambient (58);
Elevated (147)
(N/A)
Response
Measured
Tuber yield averaged
across 5 field-sites;
Tuber starch content
regressed against
[03]reportsig.
± slope with
increasing [03]
Total biomass;
Seed yield
Yield
Relative annual yield
Yield
Yield
Seed mass per plant
Seed mass per plant
Biomass
Yield
Seed protein
content;
1 ,000-seed weight
regressed across all
experiments
Fruit yield
(weight/plant)
Sugar yield
Total biomass
(g/plant)
percent change
from CF
[percent change
from ambient)
N/A (-27 % -+27%,
most comparisons
n.s.) Linear
regression slope =
-0.0098)
-25(N/A)
-13 to 20 (N/A)
From n.s. to -30
across all cultivars
N/A (2xfaster
decrease in yield/yr)
-46 (-9)
N/A
(-15 in 2002;
-25 in 2003)
-30 (N/A)
-34 (N/A)
-18 (N/A)
N/A (-17.20)
N/A (significant
negative correlation)
N/A (sig negative
correlation)
-16 (N/A)
N/A (-9)
-40 (-30)
Reference
Vandermeiren
et al. (2005)
Reid, etal.
(2008)
Sawada and
Kohno (2009)
Volketal.
(2006)
Bou Jaude
et al. (2008)
Morgan et al.
(2006)
Booker and
Fiscus (2005)
Booker etal.
(2004a)
Bulbovas et al.
(2007)
Betzelberger
et al. (2010)
Piikki et al.
(2008a)
Keutgenetal.
(2005)
De
Temmerman
et al. (2007)
Grantzand Vu
(2009)
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-------
Species
Facility
Location
Sweet Potato
Growth chambers
Bonn, Germany
Tomato
(Lycopersicon
esculentum)
OTC
Valencia, Spain
Trifolium
Subterraneum
OTC; 2.5-L pots
Madrid, Spain
Watermelon
(Citrullus lanatus)
OTC
Valencia, Spain
Yellow Nutsedge
OTC; 9-L pots
Exposure
Duration
4wk
133 days in
1998
29 days
2000, 2001 .
90 days
8wk
Os Exposure
(Additional Treatment)
8-h avg: CF (0 ppb),
Ambient (<40 ppb) Elevated
(255 ppb)
(N/A)
8-h mean ppb:
CF 16.3, NF 30.1,
NF+ 83.2
(Various cultivars; early & late
harvest)
12-havg:CF(<7.9±6.3);
Ambient (34.4±1 0.8);
Elevated (56.4±22.3)
(N: 5, 15 & 30 kg/ha)
AOT40: CF (0 ppm-h)
Ambient (5.7 ppm-h), Elevated
(34.1 ppm-h)
(N:0, 14.0 & 29.6 g/pot)
12-h avg: 12.8 ±0.6;
79.9 ±6.3; 122.7 ±9.7
(N/A)
Response
Measured
Tuber weight
Yield
Above-ground
biomass
total fruit yield (kg)
above-ground
biomass
percent change
from CF
[percent change
from ambient)
-14 (-11. 5)
n.s (n.s.)
-45 (-35)
n.s. (54)
n.s. (n.s.)
Reference
Keutgen et al.
(2008)
Calvo et al.
(2005)
Sanz et al.
(2005)
Calatayud
et al. (2006)
Grantz and
Shrestha
(2006)
In studies where variables other than 03 were included in the experimental design, response to 03 is only provided for the control level of those
variables.
Table 9-18 Summary of studies of effects of ozone exposure on growth of
natural vegetation
Species
Facility
Location
Yellow nutsedge (Cyperus
esculentus)
CSTR
Parlier, CA
35 herbaceous species
OTC
Corvallis, OR
Highbush blackberry (Rubus
argutus)
OTC
Auburn, AL
Horseweed (Conyza
canadensis)
CSTR
San Joaquin Valley, CA
Red Oak (Quercus rubrum)
Forest sites
Look Rock & Twin Creeks
Forests, TN
Exposure
Duration
53 days in
2008
1999-2002,
May-August
2004,
May-August
2005, 2 runs,
28 days each
(July-Aug,
Sept)
2001-2003,
April-October
O3 Exposure
(Additional
Treatment)
12-h mean ppb:
CF (4); CF+ (60);
CF2+(115)
4-yr avg; yearly
W1 26 ppm-h:
CF (0),
CF+ (21),
CF 2+ (49.5)
12-h mean ppb:
CF(21.7),
Ambient (32.3),
Elevated (73.3)
W126ppm-hr:
CF(0),
CF+(11),
CF 2+ (30)
(Glyphosate
resistance)
AOT60:
2001 (11.5),
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
Response
Measured
Above-ground
biomass; tubers
(g/plant)
Total community
above-ground biomass
(35 species) after 4
years
Vegetative regrowth
after pruning
Total biomass (g/plant)
Annual circumference
increment (change
relative to 2001 in year
2002;2003)
Response
ns;CF(4.1)CF+(3.9)
CF2+(2.7)
CF (459 g/m2), CF+
(457 g/m3, CF2+
(398 g/m2)
CF (75.1 g/plant),
Ambient (76.4
g/plant),
Elevated (73.1
g/plant)
Glyphosate sensitive:
CF (0.354)
CF+ (0.197)
CF2+(0.106)
Glyphosate resistant:
CF(0.510)
CF+(0.313)
CF2+(0.143)
-42.8%; +1%
Reference
Grantz etal. (201 Ob)
Pfleegeretal. (2010)
Ditchkoff etal. (2009)
Grantz etal. (2008)
McLaughlin et al.
(2007a)
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9-146
September 2011
-------
Species
Facility
Location
Pine species
Forest sites
Look Rock Forest, TN
Hickory species
Forest sites
Look Rock Forest, TN
Chestnut Oak (Quercus
prints)
Forest sites
Look Rock Forest, TN
Black Cherry (Prunus rigida)
Forest sites
Twin Creeks Forest, TN
Shortleaf pine (Pinus
echinata)
Forest sites
Twin Creeks Forest, TN
Hemlock (Tsuga
canadensis)
Forest sites
Twin Creeks Forest, TN
Red Maple (Acerrubrum)
Forest sites
Twin Creeks Forest, TN
Yellow Poplar (Liriodendron
tulipifera)
Forest sites
Look Rock, Oak Ridge, &
Twin Creeks Forest, TN
Exposure
Duration
2001-2003,
April-October
2001-2003,
April-October
2001-2003,
April-October
2002-2003,
April-October
2002-2003,
April-October
2002-2003,
April-October
2002-2003,
April-October
2002-2003,
April-October
Os Exposure
(Additional
Treatment)
AOT60:
2001 (11.5),
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2001 (11.5),
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2001 (11.5),
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
&!= ^sponse
Annual circumference
increment (change 62 5o/ . 2 9%
relative to 2001 in year '
2002;2003)
Annual circumference
increment (change 1 4% . ono,
relative to 2001 in year " ' * *' JU *
2002;2003)
Annual circumference
increment (change 440/ . .rr0,
relative to 2001 in year ^ *' °° *
2002;2003)
Annual circumference
increment (change 7j-0/
relative to 2003 in year ~'°*
2002)
Annual circumference
increment (change icao/
relative to 2003 in year •|b'b*
2002)
Annual circumference
increment (change 2i 9%
relative to 2003 in year " '
2002)
Annual circumference
increment (change r0 coi
relative to 2003 in year "OS'D*
2002)
Annual circumference
increment (change /ICQO/ icoco/
relative to 2001 in -40.y*, -IO.^OA
years 2002; 2003)
Reference
McLaughlin et al.
(2007a)
McLaughlin et al.
(2007a)
McLaughlin etal.
(2007a)
McLaughlin etal.
(2007a)
McLaughlin etal.
(2007a)
McLaughlin etal.
(2007a)
McLaughlin etal.
(2007a)
McLaughlin etal.
(2007a)
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9-147
September 2011
-------
Species
Facility
Location
Sugar Maple (Acer
saccharum)
Forest sites
Twin Creeks Forest, TN
Trembling aspen (Populus
tremuloides), 5 genotypes
Aspen FACE
Rhinelander, Wl
Hybrid Poplar (Populus
trichocarpa x Populus
deltoides)
OTC
Seattle, WA
Exposure
Duration
2002-2003,
April-October
1998-2004,
May-October
2003,
3 months
Os Exposure
(Additional
Treatment)
AOT60:
2002 (24.0),
2003(11.7)
(Observational
study with
multiple
environmental
variables)
Cumulative avg
90-day 12-h
W126.
Ambient 3.1
ppm-h Elevated:
27.2 ppm-h
(Competition with
birch, maple)
Daily mean
(ug/g):
CF(<9),
Elevated (85-128)
MeaS ResP°"se
Annual circumference
increment (change Ro H0/
relative to 2003 in year ~M-ot>
2002)
main stem volume Ambient: 6.22 dm3.
after 7 years Elevated: 4.73 dm
Total biomass ^^etevated:
Reference
Mclaughlin etal.
(2007a)
Kubiske et al. (2006)
Woo and Hinckley
(2005)
In studies where variables other than 03 were included in the experimental design, response to 03 is only provided for the control level of those
variables.
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Funct21: 707-718. http://dx.doi.ora/10.1007/s00468-007-0163-x.
Yan. K: Chen. W: He. XY: Zhang. GY: Xu. S: Wang. LL. (2010). Responses of photosynthesis, lipid peroxidation
and antioxidant system in leaves of Quercus mongolica to elevated O3. Environ Exp Bot 69: 198-204.
http://dx.doi.org/10.1016/i.envexpbot.2010.03.008.
Yoshida. S: Tamaoki. M: loki. M: Ogawa. D: Sato. Y: Aono. M: Kubo. A: Sail. S: Sail. H: Satoh. S: Nakaiima. N.
(2009). Ethylene and salicylic acid control glutathione biosynthesis in ozone-exposed Arabidopsis
thaliana. Physiol Plant 136: 284-298. http://dx.doi.Org/10.1111/i. 1399-3054.2009.01220.x.
Younglove. T: McCool. PM: Musselman. RC: Kahl. ME. (1994). Growth-stage dependent crop yield response to
ozone exposure. Environ Pollut 86: 287-295. http://dx.doi.org/10.1016/0269-7491(94)90169-4.
Yuan. JS: Himanen. SJ: Holopainen. JK: Chen. F: Stewart. CN. Jr. (2009). Smelling global climate change:
Mitigation of function for plant volatile organic compounds. Trends Ecol Evol 24: 323-331.
http://dx.doi.0rg/10.1016/i.tree.2009.01.012.
Yun. S. -C: Laurence. JA. (1999). The response of sensitive and tolerant clones of Populus tremuloides to
dynamic ozone exposure under controlled environmental conditions. New Phytol 143: 305-313.
Zak, PR: Holmes, WE: Pregitzer, KS. (2007). Atmospheric CO2 and O3 alter the flow of N15 in developing
forest ecosystems. Ecology 88: 2630-2639.
Zhang. C: Tian. HQ: Chappelka. AH: Ren. W: Chen. H: Pan. SF: Liu. ML: Stvers. DM: Chen. GS: Wang. YH.
(2007a). Impacts of climatic and atmospheric changes on carbon dynamics in the Great Smoky
Mountains National Park. Environ Pollut 149: 336-347. http://dx.doi.Org/10.1016/i.envpol.2007.05.028.
Zhang. J: Schaub. M: Ferdinand. JA: Skellv. JM: Steiner. KG: Savage. JE. (201 Oa). Leaf age affects the
responses of foliar injury and gas exchange to tropospheric ozone in Prunus serotina seedlings.
Environ Pollut 158: 2627-2634. http://dx.doi.Org/10.1016/i.envpol.2010.05.003.
Zheng. F: Wang. X: Lu. F: Hou. P: Zhang. W: Duan. X: Zhou. X: Ai. Y: Zheng. H: Ouvang. Z: Feng. Z. (2011).
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China. Global Change Biol 17: 898-910. http://dx.doi.Org/10.1111/i.1365-2486.2010.02258.x.
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10 THE ROLE OF TROPOSPHERIC OZONE IN
CLIMATE CHANGE AND UV-B EFFECTS
10.1 Introduction
1 Atmospheric O3 plays an important role in the Earth's energy budget by interacting with
2 incoming solar radiation and outgoing infrared radiation. Over mid-latitudes,
3 approximately 90% of the total atmospheric O3 column is located in the stratosphere (Kar
4 etal.. 2010; Crist etal.. 1994). Therefore, tropospheric O3 makes up a relatively small
5 portion (-10%) of the total column of O3 over mid-latitudes, but it does play an
6 important role in the overall radiation budget. The next section (Section 10.2) briefly
7 describes the physics of the earth's radiation budget, providing background material for
8 the subsequent two sections assessing how perturbations in tropospheric O3 might affect
9 (1) climate through its role as a greenhouse gas (Section 10.3), and (2) health, ecology
10 and welfare through its role in shielding the earth's surface from solar ultraviolet
11 radiation (Section 10.4).
10.2 Physics of the Earth's Radiation Budget
12 Radiant energy from the sun enters the atmosphere in a range of wavelengths, but peaks
13 strongly in the visible (400 nm up to 750 nm) part of the spectrum. Longer wavelength
14 infrared (750 nm up to ~1 mm) and shorter wavelength ultraviolet (400 nm down to
15 100 nm) radiation are also present in the solar electromagnetic spectrum. Since the
16 energy possessed by a photon is inversely proportional to its wavelength, infrared (IR)
17 radiation carries the least energy per photon, and ultraviolet (UV) radiation carries the
18 most energy per photon. UV radiation is further subdivided into classes based on
19 wavelength: UV-A refers to wavelengths from 400-315 nm; UV-B from 315-280 nm; and
20 UV-C from 280-100 nm. By the same argument above describing the relationship
21 between photon wavelength and energy, UV-A radiation is the least energetic and UV-C
22 is the most energetic band in the UV spectrum.
23 The wavelength of radiation also determines how the photons interact with the complex
24 mixture of gases, clouds, and particles present in the atmosphere (see Figure 10-1). UV-A
25 radiation can be scattered but is not absorbed to any meaningful degree by atmospheric
26 gases including O3. UV-B radiation is absorbed and scattered in part within the
27 atmosphere. UV-C is almost entirely blocked by the Earth's upper atmosphere, where it
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1
2
participates in photoionization and photodissociation processes including absorption by
stratospheric O3.
*
Backscattered
Radiation
Incident Solar UV Radiation
Stratospheric O3
Source: 2006 O3 AQCD.
Figure 10-1 Diagram of the factors that determine human exposure to
ultraviolet radiation.
4
5
6
1
8
9
10
11
12
13
Since UV-A radiation is less energetic and does not interact with O3 in the troposphere or
the stratosphere and UV-C radiation is almost entirely blocked by stratospheric O3, UV-
B radiation is the most important band to consider in relation to tropospheric O3
shielding. Furthermore, tropospheric O3 plays a "disproportionate" role in absorbing UV-
B radiation compared with stratospheric O3 on a molecule per molecule basis (Balis et
al.. 2002; Zerefos et al.. 2002; Crist etal. 1994; Bruhl and Crutzen. 1989). This effect
results from the higher atmospheric pressure present in the troposphere, resulting in
higher concentrations of gas molecules present that can absorb or scatter radiation. For
this reason, the troposphere is referred to as a "multiple scattering" regime for UV
absorption, compared to the "single scattering" regime in the stratosphere. Thus, careful
quantification of atmospheric absorbers and scatterers, along with a well-resolved
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1 description of the physics of these interactions, is necessary for predicting the impact of
2 tropospheric O3 on UV-B flux at the surface.
3 Solar flux at all wavelengths has a temporal dependence, while radiative scattering and
4 absorption have strong wavelength, path length, and gas/particle concentration
5 dependencies. These combine to create nonlinear effects on UV flux at the Earth's
6 surface. Chapter 10 of the 2006 O3 AQCDQJ.S. EPA. 2006b) describes in detail several
7 key factors that influence the spatiotemporal distribution of ground-level UV radiation
8 flux, including: (1) long-term solar activity including sunspot cycle; (2) solar rotation; (3)
9 the position of the Earth in its orbit around the sun; (4) atmospheric absorption and
10 scattering of UV radiation by gas molecules and aerosol particles; (5) absorption and
11 scattering by stratospheric and tropospheric clouds; and (6) surface albedo. The
12 efficiencies of absorption and scattering are highly dependent on the concentration of the
13 scattering medium, particle size (for aerosols and clouds), and the altitude at which these
14 processes are occurring. These properties are sensitive to meteorology, which introduces
15 additional elements of temporal dependency in ground-level UV radiation flux.
16 About 30% of incoming solar radiation is directly reflected back to space, mainly by
17 clouds or surfaces with high albedo (reflectivity), such as snow, ice, and desert sand.
18 Radiation that does penetrate to the Earth's surface and is absorbed can be re-emitted in
19 the longwave (infrared) portion of the spectrum (750 nm up to ~1 mm); the rest goes into
20 evaporating water or soil moisture or emerges as sensible heat. The troposphere is opaque
21 to the outgoing longwave radiation. Polyatomic gases such as CO2, CH4, and O3 absorb
22 and re-emit the radiation upwelling from the Earth's surface, reducing the efficiency with
23 which that energy returns to space. In effect, these gases act as a blanket warming the
24 Earth's surface. This phenomenon, known as the "Greenhouse Effect," was first
25 quantified in the 19th century (Arrhenius. 1896). and gives rise to the term "greenhouse
26 gas."
10.3 Effects of Tropospheric Ozone on Climate
Background
27 As a result of its interaction with incoming solar radiation and outgoing longwave
28 radiation, tropospheric O3 is a major greenhouse gas, and increases in its abundance may
29 contribute to climate change (IPCC. 2007b). Models estimate that the global average
30 concentration of O3 in the troposphere has doubled since the preindustrial era (Gauss et
31 al.. 2006). while observations indicate that in some regions tropospheric O3 may have
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1 increased by factors as great as 4 or 5 (Marenco et al., 1994; Staehelin et al., 1994). These
2 increases are tied to the rise in emissions of O3 precursors from human activity, mainly
3 fossil fuel consumption and agricultural processes.
4 The impact on climate of the tropospheric O3 change since preindustrial times has been
5 estimated to be about 25-40% of anthropogenic CO2 impact and about 75% of
6 anthropogenic CH4 impact (IPCC. 2007b). ranking it third in importance of the
7 greenhouse gases. In the 21st century as the Earth's population continues to grow and
8 energy technology spreads to developing countries, a further rise in the global
9 concentration of tropospheric O3 is likely, with associated consequences for human
10 health and ecosystems relating to climate change.
11 To examine the science of a changing climate and to provide balanced and rigorous
12 information to policy makers, the World Meteorological Organization (WMO) and the
13 United Nations Environment Programme (UNEP) formed the Intergovernmental Panel on
14 Climate Change (IPCC) in 1988. The IPCC supports the work of the Conference of
15 Parties (COP) to the United Nations Framework Convention on Climate Change
16 (UNFCCC). The IPCC periodically brings together climate scientists from member
17 countries of WMO and the United Nations to review knowledge of the physical climate
18 system, past and future climate change, and evidence of human-induced climate change.
19 IPCC climate assessment reports are issued every five to seven years.
20 This section draws in part on the fourth IPCC Assessment Report (AR4) (IPCC. 2007b).
21 as well as other peer-reviewed published research. Section 10.3.1 reviews evidence of
22 climate change in the recent past and projections of future climate change. It also offers a
23 brief comparison of tropospheric O3 relative to other greenhouse gases. Section 10.3.2
24 describes factors that influence the magnitude of tropospheric O3 effects on climate.
25 Section 10.3.3 considers the competing effects of O3 precursors on climate. Finally,
26 Section 10.3.4 describes the effects of changing tropospheric O3 concentrations on
27 present-day climate. Downstream effects resulting from climate change, such as
28 ecosystem responses, are outside the scope of this assessment, which focuses on the
29 direct effects of tropospheric O3 on climate.
10.3.1 Climate Change Evidence and the Influence of Tropospheric Ozone
10.3.1.1 Climate Change in the Recent Past
30 From the end of the Last Ice Age 12,000 years ago until the mid-1800s, observations
31 from ice cores show that concentrations of the long-lived greenhouse gases CO2, CH4,
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1 and N2O have been relatively stable. Unlike these greenhouse gases, O3 is not preserved
2 in ice, and no record of it before the late 1800s exists. Models, however, suggest that it,
3 too, has remained relatively constant during this time period (Thompson et al., 1993;
4 Thompson. 1992). The stable mix of greenhouse gases in the atmosphere has kept the
5 global mean temperature of the Earth close to 15°C. Without the presence of greenhouse
6 gases in the atmosphere, the Earth's temperature would be about 30°C cooler, or -15°C.
7 Since the start of the Industrial Revolution, human activity has led to significant increases
8 of greenhouse gases in the atmosphere, mainly through fossil fuel combustion. According
9 to the IPCC AR4 (IPCC. 2007b). there is now "very high confidence" that the net effect
10 of anthropogenic greenhouse gas emissions since 1750 has led to warming, and it is "very
11 likely" that human activity contributed to the 0.76°C rise in global mean temperature
12 observed over the last century. The increase of tropospheric O3 may have contributed
13 0.1-0.3°C warming to the global climate during this time period (Hansen et al.. 2005;
14 Mickley et al., 2004). Global cooling due to anthropogenic aerosols (IPCC. 2007b) has
15 likely masked the full warming effect of the anthropogenic greenhouse gases. Emissions
16 of aerosols and their precursors in the United States and other developed countries are
17 presently decreasing rapidly due to regulatory policies. The consequences of such
18 decreases on regional climate could be large, as indicated by observations (e.g., Philipona
19 et al.. 2009; Ruckstuhl et al.. 2008) and models (e.g.. Kloster et al.. 2009; Micklev et al..
20 In Press).
10.3.1.2 Projections of Future Climate Change
21 The IPCC AR4 projects a warming of ~0.2°C per decade for the remainder of the 21st
22 century (IPCC. 2007b). Even at constant concentrations of greenhouse gases in the
23 atmosphere, temperatures are expected to increase by about 0.1°C per decade, due to the
24 slow response of oceans to the warming applied so far. It is likely that the Earth will
25 experience longer and more frequent heat waves in the 21st century, together with more
26 frequent droughts and/or heavy precipitation events in some regions, due to perturbations
27 in the hydrological cycle that result from changing temperatures (IPCC. 2007b). Sea
28 levels could increase by 0.3-0.8 m by 2300 due to thermal expansion of the oceans. The
29 extent of Arctic sea ice is expected to decline, and contraction of the Greenland ice sheet
30 could further contribute to the sea level rise (IPCC. 2007b).
31 Projections of future climate change are all associated with some degree of uncertainty. A
32 major uncertainty involves future trends in the anthropogenic emissions of greenhouse
3 3 gases or their precursors. For the IPCC AR4 climate proj ections, a set of distinct
34 "storylines" or emission pathways was developed (IPCC. 2000). Each storyline took into
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1 account factors such as population growth, mix of energy technologies, and the sharing of
2 technology between developed and developing nations, and each resulted in a different
3 scenario for anthropogenic emissions. When these trends in emissions are applied to
4 models, these scenarios yield a broad range of possible climate trajectories for the 21st
5 century.
6 A second factor bringing large uncertainty to model projections of future climate is the
7 representation of climate and, especially, climate feedbacks. A rise in surface
8 temperatures would perturb a suite of other processes in the earth-atmosphere-ocean
9 system, which may in turn either amplify the temperature increase (positive feedback) or
10 diminish it (negative feedback). One important feedback involves the increase of water
11 vapor content of the atmosphere that would accompany higher temperatures (Bony et al..
12 2006). Water vapor is a potent greenhouse gas; accounting for the water vapor feedback
13 may increase the climate sensitivity to a doubling of CO2 by nearly a factor of two (Held
14 and Soden. 2000). The ice-albedo feedback is also strongly positive; a decline in snow
15 cover and sea ice extent would diminish the Earth's albedo, allowing more solar energy
16 to be deposited to the surface (Holland and Bitz. 2003; Rind etal.. 1995). A final
17 example of a climate feedback involves the effects of changing cloud cover in a warming
18 atmosphere. Models disagree on the magnitude and even the sign of this feedback on
19 surface temperatures (Soden and Held. 2006).
10.3.1.3 Metrics of Potential Climate Change
20 Two different metrics are frequently used to estimate the potential climate impact of
21 some perturbation such as a change in greenhouse gas concentration: (1) radiative
22 forcing; and (2) global warming potential (GWP).
23 Radiative forcing is a change in the radiative balance at a particular level of the
24 atmosphere or at the surface when a perturbation is introduced in the earth-atmosphere-
25 ocean system. In the global mean, radiative forcing of greenhouse gases at the tropopause
26 (top of the troposphere) is roughly proportional to the surface temperature response
27 (Hansen et al., 2005; NRC. 2005). It thus provides a useful metric for policymakers for
28 assessing the response of the earth's surface temperature to a given change in the
29 concentration of a greenhouse gas. Positive values of radiative forcing indicate warming
30 in a test case relative to the control; negative values indicate cooling. The units of
31 radiative forcing are energy flux per area, or W/m2.
32 Radiative forcing requires just a few model years to calculate, and it shows consistency
33 from model to model. However, radiative forcing does not take into account the climate
34 feedbacks that could amplify or dampen the actual surface temperature response,
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1 depending on region. Quantifying the change in surface temperature requires a climate
2 simulation in which all important feedbacks are accounted for. As these processes are not
3 well understood, the surface temperature response to a given radiative forcing is highly
4 uncertain and can vary greatly among models and even from region to region within the
5 same model.
6 GWP indicates the integrated radiative forcing over a specified period (usually 100 years)
7 from a unit mass pulse emission of a greenhouse gas or its precursor, and are reported as
8 the magnitude of this radiative forcing relative to that of CO2. GWP is most useful for
9 comparing the potential climate impacts of long-lived gases, such as N2O or CH4. Since
10 tropospheric O3 has a lifetime on the order of weeks to months, GWP is not seen as a
11 valuable metric for quantifying the importance of O3 on climate (Torster et al.. 2007).
10.3.1.4 Tropospheric Ozone as a Greenhouse Gas
12 Tropospheric O3 differs in important ways from other greenhouse gases. It is not emitted
13 directly, but is produced through photochemical oxidation of CO, CH4, and nonmethane
14 volatile organic compounds (VOCs) in the presence of nitrogen oxide radicals (NOX =
15 NO + NO2; see Section 3.2 for further details on the chemistry of O3 formation). It is also
16 supplied by vertical transport from the stratosphere. The lifetime of O3 in the troposphere
17 is typically a few weeks, resulting in an inhomogeneous distribution that varies
18 seasonally; the distribution of the long-lived greenhouse gases like CO2 and CH4 are
19 much more uniform. The longwave radiative forcing by O3 is mainly due to absorption in
20 the 9.6 urn window, where absorption by water vapor is weak. It is therefore less
21 sensitive to local humidity than the radiative forcing by CO2 or CH4, for which there is
22 much more overlap with the water absorption bands (Lenoble. 1993). And unlike other
23 major greenhouse gases, O3 absorbs in the shortwave as well as the longwave part of the
24 spectrum.
25 Figure 10-2 shows the main steps involved in the influence of tropospheric O3 on
26 climate. Emissions of O3 precursors including CO, VOCs, CH4, and NOX lead to
27 production of tropospheric O3. A change in the abundance of tropospheric O3 perturbs
28 the radiative balance of the atmosphere, an effect quantified by the radiative forcing
29 metric. The earth-atmosphere-ocean system responds to the radiative forcing with a
30 climate response, typically expressed as a change in surface temperature. Finally, the
31 climate response causes downstream climate-related health and ecosystem impacts, such
32 as redistribution of diseases or ecosystem characteristics due to temperature changes.
33 Feedbacks from both the climate response and downstream impacts can, in turn, affect
34 the abundance of tropospheric O3 and O3 precursors through multiple feedback
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1
2
3
mechanisms. Direct feedbacks are discussed further in Section 10.3.3.4; the downstream
climate impacts and their feedbacks are extremely complex and outside the scope of this
assessment.
Precursor Emissions of
CO, VOCs,CH,,NON
(Tg/y)
I
Troposphenc O,
Abundance
CIS)
I
Radiative Forcing
Due to O, Change
(WATT) "
Climate Response
fQ
I Climate Impacts ,
Figure 10-2 Schematic illustrating the effects of tropospheric ozone on climate.
Figure includes the relationship between precursor emissions,
tropospheric ozone abundance, radiative forcing, climate response,
and climate impacts. Units shown are those typical for each
quantity illustrated. Feedbacks from both the climate response and
climate impacts can, in turn, affect the abundance of tropospheric
ozone and ozone precursors through multiple feedback
mechanisms. Climate impacts are deemphasized in the figure since
these downstream effects are extremely complex and outside the
scope of this assessment.
4
5
6
7
The IPCC (2007b) reported a radiative forcing of 0.35 W/m2 for the change in
tropospheric O3 since the preindustrial era, ranking it third in importance after the
greenhouse gases CO2 (1.66 W/m2) and CH4 (0.48 W/m2). Figure 10-3 shows the global
average radiative forcing estimates and uncertainty ranges in 2005 for anthropogenic
CO2, CH4, O3 and other important agents and mechanisms. The error bars encompassing
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10-8
September 2011
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1
2
the tropospheric O3 radiative forcing estimate in the figure range from 0.25 to 0.65 W/m2,
making it relatively more uncertain than the long-lived greenhouse gases.
RADIATIVE FORCING COMPONENTS
Rr Terns
Long-lived
greenhouse gases ^
Ozone
Stratospheric water
vapour Irom CHa
Surface albedo
I Direct ettect
Cloud albedo
effect
Linear contrails
Solar irradiance
Total not
anthropogenic
RF values (W of*) Spatial scale LOSU
1.66 [1.49 to 1.83]
0.48 [0.43 to 0.53]
0.16 [0.14 to 0.18]
0.34
-0.05 [-0.15 to 0.05]
0.35 [0.25 to 0.65]
0.07 [0.02 to 0.12]
•O.2 [-0.4 to 0.0]
0.1 [0.0 to 0.2]
-0.5 [-0.9 to-0.1]
41.7 [-1.8 to -0.3]
0.01 [0.003 to 0.03]
0.12 [0.06 to 0.30]
1.6 [0.6 to 2.4]
Global
Global
Continental
to global
Global
Local to
continental
Continental
to global
Continental
to global
Continental
Global
High
High
Met)
Low
Med
- Low
Med
-Low
Low
-2 -1012
Radiative Forcing (W rtv2)
Source: Used with permission from Cambridge University Press, IPCC (2007b)
Figure 10-3 Global average radiative forcing (RF) estimates and uncertainty
ranges in 2005 for anthropogenic 062, ChU, ozone and other
important agents and mechanisms. Figure shows the typical
geographical extent (spatial scale) of the radiative forcing and the
assessed level of scientific understanding (LOSU). The net
anthropogenic radiative forcing and its range are also shown.
These require summing asymmetric uncertainty estimates from the
component terms, and cannot be obtained by simple addition.
Additional radiative forcing factors not included here are
considered to have a very low LOSU.
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10.3.2 Factors that Influence the Effect of Tropospheric Ozone on Climate
1 This section describes the main factors that influence the magnitude of the climate
2 response to changes in tropospheric O3. They include: (1) trends in the concentration of
3 tropospheric O3; (2) the effect of surface albedo on O3 radiative forcing; (3) the effect of
4 vertical distribution on O3 radiative forcing; (4) feedback factors that can alter the climate
5 response to O3 radiative forcing; and (5) the indirect effects of tropospheric O3 on the
6 carbon cycle. Trends in stratospheric O3 may also affect temperatures at the Earth's
7 surface, but aside from issues relating STE discussed in Chapter 3, Section 3.4.2,
8 stratospheric O3 assessment is beyond the scope of this document.
10.3.2.1 Trends in the Concentration of Tropospheric Ozone
9 To first order, the effect of tropospheric O3 on climate is proportional to the change in
10 tropospheric O3 concentration. The earth's surface temperatures are most sensitive to O3
11 perturbations in the mid to upper troposphere. This section therefore focuses mainly on
12 observed O3 trends in the free troposphere or in regions far from O3 sources, where a
13 change in O3 concentrations may indicate change throughout the troposphere. Data from
14 ozonesondes, mountaintops, and remote surface sites are discussed, as well as satellite
15 data.
Observed Trends in Ozone Since the Preindustrial Era
16 Measurements of O3 at two European mountain sites dating from the late 1800s to early
17 1900s show values at about 10 ppb, about one-fifth the values observed today at similar
18 sites (Pavelin etal. 1999; Marenco et al., 1994). The accuracy of these early
19 measurements is questionable however, in part because they exhibit O3 concentrations
20 equivalent to or only a couple of parts per billion greater than those observed at nearby
21 low-altitude sites during the same time period (Mickley et al.. 2001; Volz and Kiev.
22 1988). A larger vertical gradient in tropospheric O3 would be expected because of its
23 stratospheric source and its longer lifetime aloft. In another study, Staehelin et al. (1994)
24 revisited observations made in the Swiss mountains during the 1950s and found a
25 doubling in O3 concentrations from that era to 1989-1991.
26 Routine observations of O3 in the troposphere began in the 1970s with the use of balloon-
27 borne ozonesondes, but even this record is sparse. Trends from ozonesondes have been
28 highly variable and dependent on region (Logan et al.. 1999). Over most sites in the U.S.,
29 ozonesondes reveal little trend. Over Canada, observations show a decline in O3 between
30 1980 and 1990, then a rebound in the following decade (Tarasick et al.. 2005).
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1 Ozonesondes over Europe give a mixed picture. European ozonesondes showed
2 significant increases in the 1970s and 1980s, with smaller increases or even declines
3 since then (Oltmans et al., 2006; Logan et al., 1999). Over Japan, O3 in the lower
4 troposphere increased about 0.2-0.4 ppb/y during the 1990s (Naja and Akimoto. 2004).
5 Ground-based measurements in remote regions provide a record of tropospheric O3
6 extending as far back as the late 1960s or, for ship measurements, the late 1970s. A long-
7 term record of O3 in the San Bernardino Mountains of California reveals that the number
8 of high O3 days (defined as days with daily maximum O3 levels above 95 ppb) rose from
9 about 100 per summer in 1969 to over 160 in 1978 (Lee et al.. 2003a). Over the next 20
10 years, the number of high O3 days dropped slowly, to well below 100 per summer by the
11 end of the record in 1999. Springtime O3 observations from several other mountain sites
12 in the western U.S. show a positive trend of about of 0.5-0.7 ppb/y since the 1980s
13 (Cooper et al.. 2010; Jaffe et al.. 2003). Ship-borne O3 measurements for the time period
14 1977 to 2002 indicate increases of 0.1-0.7 ppb/y over the tropical and South Atlantic, but
15 no significant change over the North Atlantic (Lelieveld et al.. 2004). The lack of trend
16 for the North Atlantic would seem at odds with O3 observations at Mace Head on the
17 west coast of Ireland, which show a significant positive trend of about 0.5 ppb/y from
18 1987 to 2003 (Simmonds et al.. 2004). Over Japan, O3 at a remote mountain site has
19 increased 1 ppb/y from 1998 to 2003 (Tanimoto. 2009). a rate more than double that
20 recorded by ozonesondes in the lower troposphere over Japan during the 1990s (Najaand
21 Akimoto. 2004). At Zugspitze, a mountain site in Germany, O3 increased by 12% per
22 decade during the 1970s and 1980s, consistent with European ozonesondes (Oltmans et
23 al.. 2006). Since then, O3 continues to increase at Zugspitze, but more slowly. What little
24 data exist for the Southern Hemisphere point to significant increases in tropospheric O3
25 in recent decades, as much as -15% at Cape Grim in the 1989-2004 time period (Oltmans
26 et al.. 2006).
27 The satellite record is now approaching a length that can be useful for diagnosing trends
28 in the total tropospheric O3 column (details on the use of satellites to measure
29 tropospheric O3 are covered in Section 3.5.5.5). In contrast to the surface data from ships,
30 tropospheric O3 columns from the Total Ozone Mapping Spectrometer (TOMS) show no
31 trend over the tropical Atlantic for the period 1980-1990 (Thompson and Hudson. 1999).
32 Over the Pacific, a longer, 25 year record of TOMS data again reveals no trend over the
33 tropics, but shows increases in tropospheric column O3 of about 2-3 Dobson Units (DU)1
34 at midlatitudes in both hemispheres (Ziemke et al., 2005).
1 The Dobson Unit is a typical unit of measure for the total O3 in a vertical column above the Earth's surface. One DU is equivalent
to the amount of O3 that would exist in a 1 |jm (1CT5 m) thick layer of pure O3 at standard temperature (0°C) and pressure (1 atm),
and corresponds to a column of O3 containing 2.69 x 1020 molecules/m2. Atypical value for the amount of ozone in a column of the
Earth's atmosphere, although highly variable, is 300 DU and approximately 10% (30 DU) of that exists in the troposphere at mid
latitudes.
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1 Interpreting these recent trends in tropospheric O3 is challenging. The first difficulty is
2 reconciling apparently contradictory trends in the observations, e.g., over tropical oceans.
3 A second difficulty is that the O3 trends depend on several factors, not all of which can be
4 well characterized. These factors include (1) trends in emissions of O3 precursors, (2)
5 variation in the stratospheric source of O3, (3) changes in solar radiation resulting from
6 stratospheric O3 depletion, and (4) trends in tropospheric temperatures (Tusco and Logan.
7 2003). The trends in O3 in the San Bernardino Mountains reported by Lee et al. (2003a)
8 likely reflects regional increases in population and motor vehicles usage, and subsequent
9 implementation of more stringent motor vehicle emissions controls. More recent positive
10 trends in the western U.S. and over Japan are consistent with the rapid increase in
11 emissions of O3 precursors from mainland Asia and transport of pollution across the
12 Pacific (Cooper et al.. 2010; Tanimoto. 2009). The satellite trends over the northern mid-
13 latitudes are consistent with this picture as well (Ziemke et al.. 2005). Increases in
14 tropospheric O3 in the Southern Hemisphere are also likely due to increased
15 anthropogenic NOX emissions, especially from biomass burning (Fishman et al.. 1991).
16 Recent declines in summertime O3 over Europe can be partly explained by decreases in
17 O3 precursor emissions there (Jonson et al., 2005). while springtime increases at some
18 European sites are likely linked to changes in stratospheric dynamics (Ordonez et al..
19 2007). Over Canada, Fusco and Logan (2003) found that O3 depletion in the lowermost
20 stratosphere may have reduced the stratospheric flux of O3 into the troposphere by as
21 much as 30% from the early 1970s to the mid 1990s, consistent with the trends in
22 ozonesondes there.
Calculation of Ozone Trends for the Recent Past
23 Attempts to simulate trends in tropospheric O3 allow us to test current knowledge of O3
24 processes and to predict with greater confidence trends in future O3 concentrations.
25 Time-dependent emission inventories of O3 precursors have also been developed (for
26 1850-2000. Lamarque et al.. 2010; for 1890-1990. Van Aardenne et al.. 2001). These
27 inventories allow for the calculation of changing O3 concentration over time.
28 One recent multi-model study calculated an increase in the O3 concentration since
29 preindustrial times of 8-14 DU, or about 30-70% (Gauss et al.. 2006). The large spread in
30 modeled estimates reveals our limited knowledge of processes in the pristine atmosphere.
31 Models typically overestimate the late nineteenth and early twentieth century
32 observations available in surface air and at mountain sites by 50-100% (Lamarque et al..
33 2005: Shindell et al.. 2003: Micklev et al.. 2001: Kiehletal.. 1999). Reconciling the
34 differences between models and measurements will require more accurate simulation of
35 the natural sources of O3 (Micklev et al.. 2001) and/or implementation of novel sinks
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1 such as bromine radicals, which may reduce background O3 in the pristine atmosphere by
2 as much as 30% (Yang et al.. 2005c).
3 For the more recent past (since 1970), application of time-dependent emissions reveals an
4 equatorward shift in the distribution of tropospheric O3 in the Northern Hemisphere due
5 to the industrialization of societies at low-latitudes (Lamarque et al.. 2005; Berntsen et
6 al., 2000). By constraining a model with historical (1950s-2000) observations, Shindell et
7 al. (2002) calculated a large increase of 8.2 DU in tropospheric O3 over polluted
8 continental regions since 1950. Their result appears consistent with the large change in
9 tropospheric O3 since preindustrial times implied by the observations from the late 1800s
10 (Pavelin et al.. 1999; Marenco et al.. 1994).
10.3.2.2 The Effect of Surface Albedo on Ozone Radiative Forcing
11 The Earth's surface albedo plays a role in O3 radiative forcing. Through most of the
12 troposphere, absorption of incoming shortwave solar radiation by O3 is small relative to
13 its absorption of outgoing longwave terrestrial radiation. However, over surfaces
14 characterized by high albedo (e.g., over snow, ice, or desert sand), incoming radiation is
15 more likely to be reflected than over darker surfaces, and the probability that O3 will
16 absorb shortwave solar radiation is therefore larger. In other words, energy that would
17 otherwise return to space may instead be deposited in the atmosphere. Several studies
18 have shown that transport of O3 to the Arctic from mid-latitudes leads to radiative forcing
19 estimates greater than 1.0 W/m2 in the region, especially in summer (Shindell et al., 2006;
20 Liao et al.. 2004b: Mickley et al.. 1999). Because the Arctic is especially sensitive to
21 radiative forcing through the ice-albedo feedback, the large contribution in the shortwave
22 solar spectrum to the total radiative forcing in the region may be important.
10.3.2.3 The Effect of Vertical Distribution on Ozone Radiative
Forcing
23 In the absence of feedbacks, O3 increments near the tropopause produce the largest
24 increases in surface temperature (Lacis et al.. 1990; Wang et al.. 1980). This is a result of
25 the colder temperature of the tropopause relative to the rest of the troposphere and
26 stratosphere. Since radiation emitted by the atmosphere is approximately proportional to
27 the fourth power of its temperature2, the colder the added O3 is relative to the earth's
2 As described by the Stefan-Boltzmann law, an ideal blackbody-which the atmosphere approximates-absorbs at all wavelengths
and re-radiates proportional to the fourth power of its temperature.
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1 surface, the weaker the radiation emitted and the greater the "trapping" of longwave
2 radiation in the troposphere.
10.3.2.4 Feedback Factors that Alter the Climate Response to
Changes in Ozone Radiative Forcing
3 Estimates of radiative forcing provide a first-order assessment of the effect of
4 tropospheric O3 on climate. In the atmosphere, climate feedbacks and transport of heat
5 alter the sensitivity of Earth's surface temperature to addition of tropospheric O3.
6 Assessment of the full climate response to increases in tropospheric O3 requires use of a
7 climate model to simulate these interactions.
8 Due to its short lifetime, O3 is heterogeneously distributed through the troposphere.
9 Sharp horizontal gradients exist in the radiative forcing of O3, with the greatest radiative
10 forcing since preindustrial times occurring over the northern mid-latitudes (more on this
11 in Section 10.3.4). If climate feedbacks are particularly powerful, they may obscure or
12 even erase the correlation between regional radiative forcing and climate response
13 (Harvey. 2004; Boer and Yu. 2003). For example, several model studies have reported
14 that the horizontal pattern of surface temperature response from 2000-2100 trends in
15 predicted short-lived species (including O3) closely matches the pattern from the trends
16 in the long-lived greenhouse gases over the same time period (Levy et al., 2008; Shindell
17 et al.. 2008; Shindell et al.. 2007). This correspondence occurs even though the patterns
18 of radiative forcing for the short-lived and long-lived species differ significantly. In a
19 separate paper, Shindell (2007) found that Arctic temperatures are especially sensitive to
20 the mid-latitude radiative forcing from tropospheric O3.
21 Other studies have found that the signature of warming due to tropospheric O3 does show
22 some consistency with the O3 radiative forcing. For example, Mickley et al. (2004)
23 examined the change in O3 since preindustrial times and found greater warming in the
24 Northern Hemisphere than in the Southern Hemisphere (+0.4°C versus +0.2°C), as well
25 as higher surface temperatures downwind of Europe and Asia and over the North
26 American interior in summer. For an array of short-lived species including O3, Shindell
27 and Faluvegi (2009) found that radiative forcing applied over northern mid-latitudes yield
28 more localized responses due to local cloud, water vapor, and albedo feedbacks than
29 radiative forcing applied over the tropics.
30 Climate feedbacks can also alter the sensitivity of surface temperature to the vertical
31 distribution of tropospheric O3. The previous section (Section 10.3.2.3) described the
32 greater impact of O3 added to the upper troposphere (near the tropopause) on radiative
33 forcing, relative to additions in the mid- to lower troposphere. However, warming
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1 induced by increased O3 in the upper troposphere could stabilize the atmosphere to some
2 extent, limiting the transport of heat to the Earth's surface and mitigating the impact of
3 the added O3 on surface temperature (Joshi etal.. 2003; Christiansen. 1999). Hansen et
4 al. (1997) determined that allowing cloud feedbacks in a climate model meant that O3
5 enhancements in the mid-troposphere had the greatest effect on surface temperature.
6 Finally, climate feedbacks can amplify or diminish the climate response of one
7 greenhouse gas relative to another. For example, Mickley et al. (2004) found a greater
8 temperature response to CO2 radiative forcing than to an O3 radiative forcing of similar
9 global mean magnitude, due in part to the relatively weak ice-albedo feedback for O3.
10 Since CO2 absorbs in the same bands as water vapor, CO2 radiative forcing saturates in
11 the middle troposphere and is also shifted toward the drier poles. A poleward shift in
12 radiative forcing amplifies the ice-albedo feedback in the case of CO2, and the greater
13 mid-troposphere radiative forcing allows for greater surface temperature response,
14 relative to that for O3.
10.3.2.5 Indirect Effects of Tropospheric Ozone on the Carbon Cycle
15 A proposed indirect effect of tropospheric O3 on climate involves the carbon cycle. By
16 directly damaging plant life in ways discussed in Chapter 9, increases in tropospheric O3
17 may depress the land-carbon sink of CO2, leading to accumulation of CO2 in the
18 atmosphere and ultimately warming of the Earth's surface. Sitch et al. (2007) calculated
19 that this indirect warming effect of O3 on climate has about the same magnitude as the
20 O3 direct effect. Their results suggest a doubled sensitivity of surface temperatures to O3
21 radiative forcing, compared to current model estimates.
10.3.3 Competing Effects of Ozone Precursors on Climate
22 Changes in O3 precursors affect not just O3 concentrations, but also other species that
23 have importance to the radiative balance of the earth's climate system. More specifically,
24 O3 and its precursors exert a strong control on the oxidizing capacity of the troposphere
25 (Derwent et al.. 2001). For example, an increase in CO or VOCs would lead to a decrease
26 in hydroxyl (OH) concentrations. Since OH is a major sink of the greenhouse gas CH4, a
27 decline in OH would lengthen the CH4 lifetime, enhance the CH4 concentration, and
28 amplify surface warming. A rise in NOX emissions, on the other hand, could lead to an
29 increase in OH in certain locations, shortening the CH4 lifetime and leading to surface
30 cooling (Fuglestvedt et al.. 1999). O3 can itself generate OH through (1) photolysis
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1 leading to excited oxygen atoms followed by reaction with water vapor and (2) reaction
2 withHO2.
3 Analyzing the net radiative forcing per unit emission for a suite of O3 precursors,
4 Shindell and Faluvegi (2009) calculated positive (+0.25 W/m2) radiative forcing from the
5 increase in anthropogenic emissions of CO and VOCs since preindustrial times, as well
6 as for CH4 (+1 W/m2). These species also contribute to warming via their eventual
7 contribution to CO2. In contrast, Shindell and Faluvegi (2009) found negative (-
8 0.29 W/m2) radiative forcing from anthropogenic emissions of NOX due mainly to the
9 link between NOX and CH4. These results are consistent with those of Forster et al.
10 (2007) who reported a net warming of+0.27 W/m2 for combined anthropogenic CO and
11 VOCs emissions and a net cooling of -0.21 W/m2 for anthropogenic NOX emissions.
12 Other studies have found a near cancellation of the positive O3 radiative forcing and the
13 negative CH4 radiative forcing that arise from an incremental increase in anthropogenic
14 NOX emissions (Naiket al.. 2005; Fiore et al.. 2002; Fuglestvedt et al.. 1999).
15 The net effect of aircraft NOX on climate is complex. While Isaksen et al. (2001) reported
16 that the net radiative forcing effect of aircraft NO emissions is near zero, Wild et al.
17 (2001) calculated a net warming due to increased O3 production efficiency in the upper
18 troposphere. More recently, Stevenson et al. (2004) completed a detailed analysis of the
19 OH budget in the years following a pulse of aircraft NOX emissions. They calculated that
20 while such a pulse leads initially to warming through O3 production over a few months,
21 the long-term effect is cooling through the effects on CH4. Both aircraft NOX and the O3
22 it generates enhance OH concentrations, with the longer-lived O3 responsible for
23 transferring the oxidizing effects of aircraft emissions away from flight corridors.
24 Finally, OH production from O3 precursors can affect regional sulfate air quality and
25 climate forcing by increasing gas-phase oxidation rates of SO2. Using the A1B scenario
26 in the IPCC AR4, Unger et al. (2006) reported that at 2030, enhanced OH from the A IB
27 O3 precursors increased surface sulfate aerosol concentrations by up to 20% over India
28 and China, relative to the present-day, with a corresponding increase in radiative cooling
29 over these regions. In this way, O3 precursors may impose an indirect cooling via sulfate
30 (Unger. 2006).
31 Taken together, these results point out the need for careful assessment of net radiative
32 forcing involving multiple pollutants in developing climate change policy (Unger et al.,
3 3 2008). Naik et al. (2005) calculated that a carefully combined reduction of CO, VOCs,
34 and NOX emissions could lead to net cooling, especially over the tropics. Several studies
3 5 point to CH4 as a particularly attractive target for emissions control since CH4 is itself an
36 important precursor of O3 (West et al.. 2007; Fiore et al.. 2002). Shindell et al. (2005)
37 calculated that the emissions-based radiative forcing of anthropogenic CH4, which
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1 includes both its own radiative forcing and that of CH4-generated O3, is 0.8-0.9 W/m2,
2 about double that of the CH4 abundance-based radiative forcing. Fiore et al. (2002) found
3 that reducing anthropogenic CH4 emissions by 50% would lead to a global negative (-
4 0.37 W/m2) radiative forcing, mostly from CH4. In later research, Fiore et al. (2008)
5 reported that CH4 reductions would most strongly affect tropospheric O3 column
6 amounts in a zonal band centered around 30 N, a region of strong downwelling and NOX-
7 saturated conditions near the surface.
10.3.4 Calculating Radiative Forcing and Climate Response to Past Trends
in Tropospheric Ozone
8 The magnitude of the radiative forcing from the change in tropospheric O3 since the
9 preindustrial era is uncertain. This uncertainty derives in part from the scarcity of early
10 measurements and in part from our limited knowledge regarding processes in the natural
11 atmosphere. As noted previously, the IPCC AR4 reports a radiative forcing of 0.35 W/m2
12 from the change in tropospheric O3 since 1750 (Torster et al.. 2007). ranking it third in
13 importance among greenhouse gases after CO2 and CH4. The O3 radiative forcing could,
14 in fact, be as large as 0.7 W/m2, if reconstructions of preindustrial and mid-20th century
15 O3 based on the measurement record are valid (Shindell and Faluvegi. 2002; Mickley et
16 al.. 2001). In any event, Unger et al. (2010) showed that present-day O3 radiative forcing
17 can be attributed to emissions from many economic sectors, including on-road vehicles,
18 household biofuel, power generation, and biomass burning. As much as one-third of the
19 radiative forcing from the 1890 to 1990 change in tropospheric O3 could be due to
20 increased biomass burning (Ito etal. 2007a).
21 These calculated radiative forcing estimates can be compared to those obtained from
22 satellite data. Using data from TOMS, Worden et al. (2008) estimated a reduction in
23 clear-sky outgoing longwave radiation of 0.48 W/m2 by O3 in the upper troposphere over
24 oceans in 2006. This radiative forcing includes contributions from both anthropogenic
25 and natural O3. Assuming that the concentration of O3 has roughly doubled since
26 preindustrial times (Gauss et al.. 2006). the total O3 radiative forcing estimated with
27 TOMS is consistent with that obtained from models estimating just the anthropogenic
28 contribution.
29 Calculation of the climate response to the O3 radiative forcing is challenging due to
30 complexity of feedbacks, as mentioned in Sections 10.3.1.2 and 10.3.2.4. In their
31 modeling study, Mickley et al. (2004) reported a global mean increase of 0.28°C since
32 preindustrial times, with values as large as 0.8°C in continental interiors. For the time
33 period since 1870, Hansen et al. (2005) estimated a much smaller increase in global mean
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1 surface temperature (0.11°C), but they implemented 1880s anthropogenic emissions in
2 their base simulation and also took into account trends in both stratospheric and
3 tropospheric O3; the modeled decline of lower stratospheric O3, especially over polar
4 regions, cooled surface temperatures in this study, counteracting the warming effect of
5 increasing tropospheric O3.
6 Figure 10-4 shows the Hansen et al. (2005) results as reported in Shindell et al. (2006). In
7 that figure, summertime O3 has the largest radiative impact over the continental interiors
8 of the Northern Hemisphere. Shindell et al. (2006) estimated that the change in
9 tropospheric O3 over the 20th century could have contributed about 0.3°C to annual mean
10 Arctic warming and as much as 0.4-0.5°C during winter and spring. Over eastern China,
11 Chang et al. (2009) calculated a surface temperature increase of 0.4°C to the 1970-2000
12 change in tropospheric O3. It is not clear, however, to what degree regional changes in
13 O3 concentration influenced this response, as opposed to more global changes.
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Annual surface air temperature
Annual radiative forcing
.41
-1.1 -.9 -.7 -J5 -.3 -.1 .1 .3 .5 .7 .9 1.1 0 .1 .2 .3 .4 .5 .6 .7
Summer (JJA) surface air temperature .10 Winter (DJF) surface air temperature
.11
-1.1 -.9 -.7 -.5 -.3 -.1 .1 .3 .5 .7 .9 1.5 -1.1 -.9 -.7 %5 -.3 -.1 .1 .3 .5 .7 .9 1.4
Source: Used with permission from American Geophysical Union (Shindell et al.. 2006)
Figure includes the input radiative forcing (W/m2), as computed by the NASA GISS chemistry-climate model. Values are surface
temperature trends for the annual average (top left), June-August (bottom left), and December-February (bottom right) and annual
average tropopause instantaneous radiative forcing from 1880 to 1990 (top right). Temperature trends greater than about 0.1 °C are
significant over the oceans, while values greater than 0.3°C are typically significant over land, except for northern middle and high
latitudes during winter where values in excess of about 0.5°C are significant. Values in the top right corner give area-weighted global
averages in the same units as the plots.
Figure 10-4 Ensemble average 1900-2000 surface temperature trends (°C per
century) in response to tropospheric ozone changes.
10.4 UV-B Related Effects and Tropospheric Ozone
i
2
3
4
5
10.4.1 Background
UV radiation emitted from the Sun contains sufficient energy when it reaches the Earth to
break (photolyze) chemical bonds in molecules, thereby leading to damaging effects on
living organisms and materials. Atmospheric O3 plays a crucial role in reducing exposure
to solar UV radiation at the Earth's surface. Stratospheric O3 is responsible for the
majority of this shielding effect, as approximately 90% of total atmospheric O3 is located
there over mid-latitudes (Kar et al., 2010; Crist et al.. 1994). Investigation of the
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September 2011
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1 supplemental shielding of UV-B radiation provided by tropospheric O3 is necessary for
2 quantifying UV-B exposure and the incidence of related human health effects, ecosystem
3 effects, and materials damage. The role of tropospheric O3 in shielding of UV-B radiation
4 is discussed in this section.
10.4.2 Human Exposure and Susceptibility to Ultraviolet Radiation
5 The factors that potentially influence UV radiation exposure were discussed in detail in
6 Chapter 10 of the 2006 O3 AQCD and are summarized here. These factors included
7 outdoor activity, occupation, age, gender, geography, and protective behavior. Outdoor
8 activity and occupation both influenced the amount of time people spend outdoors during
9 daylight hours, the predominant factor for exposure to solar UV radiation. Participation in
10 outdoor sports (e.g., basketball, soccer, golf, swimming, cycling) significantly increased
11 UV radiation exposure (Thieden et al.. 2004a: Thieden et al.. 2004b: Moehrle. 2001;
12 Moehrle et al.. 2000). Occupations that substantially increased exposure to UV radiation
13 included farming (Schenker et al.. 2002; Airey et al.. 1997). fishing (Rosenthal et al..
14 1988). landscaping (Rosenthal et al.. 1988). construction (Gies and Wright. 2003).
15 physical education (Vishvakarman et al.. 2001). mail delivery (Vishvakarman et al..
16 2001). and various other occupations that require workers to spend the majority of their
17 day outdoors during peak UV radiation hours.
18 Age and gender were found to be factors that influence human exposure to UV radiation,
19 particularly by influencing other factors of exposure such as outdoor activity and risk
20 behavior. Studies indicated that females generally spent less time outdoors and,
21 consequently, had lower UV radiation exposure compared to males (Godar etal.. 2001;
22 Gies et al.. 1998; Shoveller et al.. 1998). The lowest exposure to UV radiation among
23 Americans in the Godar et al. (2001) study was received in females during their child
24 raising years (age 22-40 years); the highest exposure was observed in males aged
25 41-59 years. A similar Canadian survey found that younger adult males had the greatest
26 exposures to UV radiation (Shoveller et al.. 1998).
27 Geography influences the degree of solar UV flux to the surface, and hence exposure to
28 UV radiation. In the U.S. study by Godar et al. (2001). northerners and southerners were
29 found to spend an equal amount of time outdoors; however, the higher solar flux at lower
30 latitudes significantly increased the annual UV radiation dose for southerners. The annual
31 UV radiation doses in southerners were 25 and 40% higher in females and males,
32 respectively, compared to northerners. Other studies also have shown that altitude and
33 latitude influence personal exposure to UV radiation (Rigel etal.. 1999; Kimlin et al..
34 1998).
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1 Protective behaviors such as using sunscreen (Nole and Johnson. 2004). wearing
2 protective clothing (Rosenthal et al.. 1988). and spending time in shaded areas (TVIoise et
3 al.. 1999) were shown to reduce exposure to UV radiation. In one study, the use of
4 sunscreen was associated with extended intentional UV radiation exposure (Autier et al..
5 1999); however, a follow-up study indicated that sunscreen use increased duration of
6 exposures to doses of UV radiation that were below the threshold level for erythema
7 (Autier etal.. 2000).
8 Given these and other factors that potentially influence UV radiation exposure, the 2006
9 O3 AQCD listed the following subpopulations potentially at risk for higher exposures to
10 UV radiation:
11 • Individuals who engage in high-risk behavior (e.g., sunbathing);
12 • Individuals who participate in outdoor sports and activities;
13 • Individuals who work outdoors with inadequate shade (e.g., farmers,
14 construction workers, etc.); and
15 • Individuals living in geographic areas with higher solar flux including lower
16 latitudes (e.g., Honolulu, HI) and higher altitudes (e.g., Denver, CO).
17 The risks associated with all these factors are, of course, highly dependent on season and
18 region (Sliney and Wengraitis. 2006).
10.4.3 Human Health Effects due to UV-B Radiation
19 Chapter 10 of the 2006 O3 AQCD covered in detail the human health effects associated
20 with solar UV-B radiation exposure. These effects include erythema, skin cancer, ocular
21 damage, and immune system suppression. These adverse effects, along with protective
22 effects of UV radiation through increased production of vitamin D are summarized in this
23 section. For additional details, the reader is referred to Chapter 10 of the 2006 O3 AQCD
24 (U.S. EPA. 2006b) and references therein.
25 The most conspicuous and well-recognized acute response to UV radiation is erythema,
26 or the reddening of the skin. Erythema is likely caused by direct damage to DNA by UV
27 radiation (Matsumura and Ananthaswamy. 2004). Many studies discussed in the 2006 O3
28 AQCD found skin type to be a significant risk factor for erythema. Additional risk factors
29 include atopic dermatitis (ten Berge et al.. 2009).
30 Skin cancer is another prevalent health effect associated with UV radiation. Exposure to
31 UV radiation is considered to be a major risk factor for all forms of skin cancer (Diepgen
32 and Mahler. 2002; Gloster and Brodland. 1996). Ultraviolet radiation is especially
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1 effective in inducing genetic mutations and acts as both a tumor initiator and promoter.
2 Keratinocytes have evolved DNA repair mechanisms to correct the damage induced by
3 UV; however, mutations can occur, leading to skin cancers that are appearing with
4 increasing frequency (Hildesheim and Fornace. 2004). The relationship between skin
5 cancer and chronic exposure to UV radiation is further explored in Chapter 10 of the
6 2006 O3 AQCD (U.S. EPA. 2006b).
7 Ocular damage from UV radiation exposure includes effects on the cornea, lens, iris, and
8 associated epithelial and conjunctival tissues. The region of the eye affected by exposure
9 to UV radiation depends on the wavelength of the incident UV radiation. Depending on
10 wavelength, common health effects associated with UV radiation include photokeratitis
11 (snow blindness; short wavelengths) and cataracts (opacity of the lens; long
12 wavelengths).
13 Experimental studies have shown that exposure to UV radiation may suppress local and
14 systemic immune responses to a variety of antigens (Clydesdale et al.. 2001; Garssen and
15 Van Loveren. 2001; Selgrade et al., 1997). In rodent models, these effects have been
16 shown to worsen the course and outcome of some infectious diseases and cancers
17 (Granstein and Matsui. 2004; Norval et al., 1999). Results from human clinical studies
18 suggest that immune suppression induced by UV radiation may be a risk factor
19 contributing to skin cancer induction (Ullrich. 2005; Caforio et al.. 2000; Lindelof et al..
20 2000). There is also evidence that UV radiation has indirect involvement in viral
21 oncogenesis through the human papillomavirus (Pfister. 2003). dermatomyositis (Okada
22 et al.. 2003). human immunodeficiency virus (Breuer-McHam et al.. 2001) and other
23 forms of immunosuppression (Selgrade et al.. 2001).
24 A potential health benefit of increased UV-B exposure relates to the production of
25 vitamin D in humans. Most humans depend on sun exposure to satisfy their requirements
26 for vitamin D (Holick. 2004). Vitamin D deficiency can cause metabolic bone disease
27 among children and adults, and also may increase the risk of many common chronic
28 diseases, including type I diabetes mellitus and rheumatoid arthritis (Holick. 2004).
29 Substantial in vitro and toxicological evidence also support a role for vitamin D activity
30 against the incidence or progression of various forms of cancer (Giovannucci. 2005; John
31 etal.. 2005; Smedbv et al.. 2005; Grant and Garland. 2004; Hughes et al.. 2004;
32 Freedman et al.. 2002: Grant. 2002a. b; John etal.. 1999: Studzinski and Moore. 1995:
33 Lefkowitz and Garland. 1994: Hanchette and Schwartz. 1992: Garland et al.. 1990:
34 Gorham et al.. 1990). In some studies, UV-B related production of vitamin D had
3 5 potential beneficial immunomodulatory effects on multiple sclerosis, insulin-dependent
36 diabetes mellitus, and rheumatoid arthritis (Ponsonby et al.. 2002: Cantorna. 2000). More
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1 details on UV-B protective studies are provided in Chapter 10 of the 2006 O3 AQCD
2 (U.S. EPA. 2006b).
3 In establishing guidelines on limits of exposure to UV radiation, the International
4 commission on Non-Ionizing Radiation Protection (ICNIRP) agreed that some low-level
5 exposure to UV radiation has health benefits (ICNIRP. 2004). However, the adverse
6 health effects of higher UV exposures necessitated the development of exposure limits
7 for UV radiation. The ICNIRP recognized the challenge in establishing exposure limits
8 that would achieve a realistic balance between beneficial and adverse health effects. As
9 concluded by ICNIRP (2004). "[t]he present understanding of injury mechanisms and
10 long-term effects of exposure to [UV radiation] is incomplete, and awaits further
11 research."
10.4.4 Ecosystem and Materials Damage Effects Due to UV-B Radiation
12 A 2009 progress report on the environmental effects of O3 depletion from the UNEP,
13 Environmental Effects Assessment Panel (UNEP. 2009) lists many ecosystem and
14 materials damage effects from UV-B radiation. An in-depth assessment of the global
15 ecosystem and materials damage effects from UV-B radiation per se is out of the scope of
16 this assessment. However, a brief summary of some mid-latitude effects is provided in
17 this section to provide context for UV-B related issues pertaining to tropospheric O3. The
18 reader is referred to the UNEP report (UNEP. 2009) and references therein for further
19 details. All of these UV-B related ecosystem and materials effects can also be influenced
20 by climate change through temperature and other meteorological alterations, making
21 quantifiable predictions of UV-B effects difficult.
22 Terrestrial ecosystem effects from increased UV-B radiation include reduced plant
23 productivity and plant cover, changes in biodiversity, susceptibility to infection, and
24 increases in natural UV protective responses. In general, however, these effects are small
25 for moderate UV-B increases at mid-latitudes. A field study on wheat in southern Chile
26 found no substantial changes in crop yield with moderate increases in UV-B radiation
27 (Calderini et al.. 2008). Similarly, field studies on silver birch (Betulapendula) in
28 Finland found no significant effects in photosynthetic function with increases in UV-B
29 radiation (Aphalo et al.. 2009). Subtle, but important, changes in habitat and biodiversity
30 have also been linked to increases in UV-B radiation (Mazza et al.. 2010; Obaraet al..
31 2008; Wahl. 2008). Some plants have natural coping mechanisms for dealing with
32 changes in UV-B radiation (Favory et al.. 2009; Jenkins. 2009; Brown and Jenkins. 2008;
33 loki et al.. 2008). but these defenses may have costs in terms of reduced growth (Snell et
34 al.. 2009: Clarke and Robinson. 2008: Semerdiieva et al.. 2003: Phoenix et al.. 2000).
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1 Aquatic ecosystem effects from increased UV-B radiation include sensitivity in
2 growth, immune response, and behavioral patterns of aquatic organisms. One study
3 looking at coccolithophores, an abundant phytoplankton group, found a 25% reduction in
4 cellular growth with UV-B exposure (Gao et al.. 2009a). Exposure to relevant levels of
5 UV-B radiation has been shown to modify immune response, blood chemistry, and
6 behavior in certain species of fish (Markkula et al.. 2009; Holtby and Bothwell. 2008;
7 Jokinen et al.. 2008). Adverse effects on growth and development from UV-B radiation
8 have also been observed for amphibians, sea urchins, mollusks, corals, and zooplankton
9 (Garcia etal.. 2009; Romansic et al.. 2009; Croteau et al.. 2008a: Croteau et al.. 2008b:
10 Marquis et al.. 2008; Marquis and Miaud. 2008; Oromi et al.. 2008). Increases in the flux
11 of UV-B radiation may also result in an increase in the catalysis of trace metals including
12 mercury, particularly in clear oligotrophic lakes with low levels of dissolved organic
13 carbon to stop the penetration of UV-B radiation (Schindler et al.. 1996). This could then
14 alter the mobility of trace metals including the potential for increased mercury
15 volatilization and transport within and among ecosystems.
16 Biogeochemical cycles, particularly the carbon cycle, can also be influenced by
17 increased UV-B radiation. A study on high latitude wetlands found UV-induced increases
18 in CO2 uptake through soil respiration (Haapala et al.. 2009) while studies on arid
19 terrestrial ecosystems found evidence for UV-induced release of CO2 through
20 photodegradation of above-ground plant litter (Brandt et al.. 2009; Henry et al.. 2008;
21 Caldwell et al.. 2007; Zepp et al., 2007). Changes in solar UV radiation may also have
22 effects on carbon cycling and CO2 uptake in the oceans (Brewer and Peltzer. 2009;
23 Meador et al.. 2009; Fritz et al.. 2008; Zepp et al.. 2008; Hader et al.. 2007) as well as
24 release of dissolved organic matter from sediment and algae (Mayer et al.. 2009;
25 Riggsbee et al.. 2008). Additional studies showing effects on these and additional
26 biogeochemical cycles including the water cycle and halocarbon cycle can be found in
27 the UNEP report (UNEP. 2009) and references therein.
28 Materials damage from increased UV-B radiation include UV-induced
29 photodegradation of wood (Kataoka et al.. 2007) and plastics (Pickett et al.. 2008). These
30 studies and others summarizing photo-resistant coatings and materials designed to reduce
31 photodegradation of materials are summarized in the UNEP report (UNEP. 2009) and
32 references therein.
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10.4.5 UV-B Related Effects Associated with Changes in Tropospheric
Ozone Concentrations
1 There are multiple complexities in attempting to quantify the relationship between
2 changes in tropospheric O3 concentrations and UV radiation exposure. Quantifying the
3 relationship between UV radiation and health or welfare effects is complicated by the
4 uncertainties involved in the selection of an action spectrum and appropriate
5 characterization of dose (e.g., peak or cumulative levels of exposure, timing of exposures,
6 etc.) The lack of published studies that critically examine these issues together-that is the
7 incremental health or welfare effects attributable specifically to UV-B changes resulting
8 from reductions in tropospheric O3 concentrations—reflects the significant challenges in
9 this field.
10 As reported in the 2006 O3 AQCD, one analysis by Lutter and Wolz (1997) attempted to
11 estimate the effects of a nationwide 10 ppb reduction in seasonal average tropospheric O3
12 on the incidence of nonmelanoma and melanoma skin cancers and cataracts in humans.
13 Their estimate, however, depended upon several simplifying assumptions, ranging from
14 an assumed generalized 10-ppb reduction in O3 column density, national annual average
15 incidence rates for the two types of skin cancer, and simple, linear biological
16 amplification factors. Specifically, the decrease of 10 ppbv in seasonally averaged O3
17 concentrations is likely an overestimate since it doesn't account for the influence of
18 background O3 coming from the global accumulation or generation of regional chemistry
19 (Adamowicz et al.. 2004). Further, the methodologies used in this analysis have ignored
20 area-specific factors that are important in estimating the extent to which small, variable
21 changes in ground-level O3 mediate long-term exposures to UV-B radiation.
22 A more recent study by Madronich et al. (2011) used CMAQ to estimate UV radiation
23 response to changes in tropospheric O3 under different control scenarios projected out to
24 2020. This study focused on southeastern U.S. and accounted for spatial and temporal
25 variation in tropospheric O3 reductions, an important consideration since most controls
26 are focused on reducing O3 in populated urban areas. The contrasting control strategies
27 considered in this study included a historical scenario designed to meet an 84 ppb 8-h
28 daily max standard and a reduced scenario designed to bring areas predicted to exceed a
29 similarly designed 70 ppb standard into attainment. A biologically effective irradiance
30 was estimated by multiplying the modeled UV irradiance by a sensitivity function (action
31 spectrum) for the induction of nonmelanoma skin cancer in mice corrected for human
32 skin transmission, then integrating over UV wavelengths. The average relative change in
33 skin cancer-weighted surface UV radiation between the two scenarios was about 0.11%
34 over June, July and August. Weighting by population, this estimate increased to 0.19%.
3 5 Madronich et al. (2011) report that their estimated UV radiation increment is an order of
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1 magnitude less than that by Lutter and Wolz (1997) with the main reason for the
2 discrepancy coming from the unrealistic uniform 10 ppb reduction in O3 assumed in the
3 former study. Madronich et al. (2011) did not attempt to link their predicted increase in
4 UV radiation to a predicted increase in skin cancer incidence, however, due to several
5 remaining and substantial uncertainties.
6 A handful of additional studies have addressed the relationship between changes in
7 tropospheric pollutant concentrations and UV-B radiation exposure, providing some
8 additional insight. A study by Palancar and Toselli (2002) looked at changes in measured
9 UV-B radiation in relation to ground-level air pollutants during several air pollution
10 episodes in Cordoba, Argentina. They found that changes in aerosol concentrations
11 explained the majority of UV-B radiation fluctuations, and that changes in tropospheric
12 O3 and SO2 had little effect. Repapis et al. (1998) performed a similar study on UV-B
13 exposures during high and low air pollution days in Athens, Greece. They found cloud
14 cover and aerosols to be the major factors in observed UV-B exposures reductions.
15 Studies by Acosta and Evans (2000) in Mexico City and Koronakis et al. (2002) in
16 Athens, Greece both found significant reductions in surface-level UV exposures during
17 pollution episodes. Both these studies include tropospheric O3 as a potential driver for the
18 reductions, but neither study was able to quantify the influence of individual atmospheric
19 components involved in the observed attenuation in UV-B radiation.
20 In the absence of reliable studies specifically addressing UV-B related health effects from
21 a reduction in tropospheric O3, inferences were made in the 2006 O3 AQCD on the basis
22 of studies focused on stratospheric O3 depletion. Studies included in that review
23 examined the potential effect of stratospheric O3 depletion on the risk of erythema
24 (Longstreth et al.. 1998). skin cancer (Urbach. 1997; Slaperetal.. 1996; De Gruiil 1995;
25 Longstreth et al.. 1995; Madronich and De Gruijl. 1993). nonmelanoma skin cancer
26 (Slaperetal.. 1996; Longstreth et al.. 1995). and cataracts (Longstreth et al.. 1995). Note
27 that several of the concerns expressed above in relation to the Lutter and Wolz (1997)
28 analysis are relevant to these analyses as well. Furthermore, these studies have a high
29 degree of uncertainty due to inadequate information on the action spectrum and dose-
30 response relationships. As a result, caution is advised when assessing and interpreting the
31 quantitative results of health risks due to stratospheric O3 depletion in the context of
32 tropospheric O3 shielding.
33 Although the UV-B related health effects attributed to marginal reductions in
34 tropospheric or ground-level O3 that would result from reductions in O3 concentrations
3 5 have not been directly assessed, they would be expected to be small given the above
36 findings and the fact that tropospheric O3 makes up only -10% of the total atmospheric
37 O3 column at mid-latitudes (Kar et al.. 2010). Furthermore, O3 present in the planetary
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1 boundary layer makes up only -10% of tropospheric O3 (Thompson et al., 2007) and the
2 NAAQS has only a fractional influence on those ground-level O3 concentrations. The net
3 result is a very small influence on total column O3 through attainment of the O3 standard.
4 In addition, the health benefits of UV-B in the production of vitamin D suggests that
5 increased risks of human disease due to a slight excess in UV-B radiation exposure may
6 be offset by the benefits of enhanced vitamin D production. However, as with other
7 impacts of UV-B on human health, this beneficial effect of UV-B has not been studied in
8 sufficient detail to allow for a credible health benefits assessment. Hence, the above
9 mentioned health and welfare effects associated with UV-B exposures resulting from
10 changes in ground-level O3 concentrations would likely be small or nonexistent based on
11 current information.
12 More reasonable estimates of the human health impacts of enhanced UV-B penetration
13 following reduced ground-level O3 concentrations require both (a) a solid understanding
14 of the multiple factors that define the extent of human exposure to UV-B, and (b) well-
15 defined and quantifiable links between human disease and UV-B exposure. Within the
16 uncertain context of presently available information on UV-B surface fluxes, a risk
17 assessment of UV-B-related health effects would need to factor in human habits (e.g.,
18 daily activities, recreation, dress, and skin care) in order to adequately estimate UV-B
19 exposure levels. Little is known about the impact of variability in these human factors on
20 individual exposure to UV radiation. Furthermore, detailed information does not exist
21 regarding the relevant type (e.g., peak or cumulative) and time period (e.g., childhood,
22 lifetime, or current) of exposure, wavelength dependency of biological responses, and
23 inter-individual variability in UV resistance. In conclusion, the effect of changes in
24 surface-level O3 concentrations on UV-induced health outcomes cannot yet be critically
25 assessed within reasonable uncertainty. The reader is referred to the U.S. EPA 2002 Final
26 Response to Court Remand (U.S. EPA. 2003) for detailed discussions of the data and
27 scientific issues associated with the determination of public health benefits resulting from
28 the attenuation of UV-B by surface-level O3.
10.5 Summary
10.5.1 Summary of the Effects of Tropospheric Ozone on Climate
29 Tropospheric O3 is a major greenhouse gas, third in importance after CO2 and CH4.
30 While the developed world has successfully reduced emissions of O3 precursors in recent
31 decades, many developing countries have experienced large increases in precursor
32 emissions and these trends are expected to continue, at least in the near term. Projections
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1 of radiative forcing due to changing O3 over the 21st century show wide variation, due in
2 large part to the uncertainty of future emissions of source gases. In the near-term (2000-
3 2030), projections of O3 radiative forcing range from near zero to +0.3 W/m2, depending
4 on the emissions scenario (Stevenson et al.. 2006). Reduction of tropospheric O3
5 concentrations could therefore provide an important means to slow climate change in
6 addition to the added benefit improving surface air quality.
7 It is clear that increases in tropospheric O3 lead to warming. However the precursors of
8 O3 also have competing effects on the greenhouse gas CH4, complicating emissions
9 reduction strategies. A decrease in CO or VOC emissions would enhance OH
10 concentrations, shortening the lifetime of CH4, while a decrease in NOX emissions could
11 depress OH concentrations in certain regions and lengthen the CH4 lifetime. Recent
12 research, however, has indicated that a carefully combined reduction of CO, VOCs, and
13 NOX emissions could lead to net cooling (Naik et al.. 2005). They calculate that such
14 reductions would have the greatest impact for developing countries in tropical regions.
15 Abatement of CH4 emissions would likely provide the most straightforward means to
16 address climate change since CH4 is itself an important precursor of background O3
17 (West et al.. 2007; West et al.. 2006; Fiore et al.. 2002). A reduction of CH4 emissions
18 would also improve air quality on its own right. A set of global abatement measures
19 identified by West and Fiore (2005) could reduce CH4 emissions by 10% at a cost
20 savings, decrease background O3 by about 1 ppb in the Northern Hemisphere summer,
21 and lead to a global net cooling of 0.12 W/m2. Unlike measures to reduce NOX, which
22 would have immediate impacts on surface O3 but little net radiative forcing, the cooling
23 effects of CH4 controls would be realized gradually, over -12 years. West et al. (2007)
24 explored further the benefits of CH4 abatement, finding that a 20% reduction in global
25 CH4 emissions would lead to significantly greater cooling per unit reduction in surface
26 O3, compared to 20% reductions in VOCs or CO.
27 Important uncertainties remain regarding the impact of tropospheric O3 on future climate
28 change. To address these uncertainties, further research is needed to: (1) enhance our
29 knowledge of the natural atmosphere; (2) interpret observed trends of O3 in the free
30 troposphere and remote regions; (3) improve our understanding of the CH4 budget,
31 especially emissions from wetlands and agricultural sources, (4) understand the
32 relationship between regional O3 radiative forcing and regional climate change; and (5)
33 determine the optimal mix of emissions reductions that would act to limit future climate
34 change.
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10.5.2 Summary of UV-B Related Effects on Human Health, Ecosystems,
and Materials Relating to Changes in Tropospheric Ozone
Concentrations
1 UV radiation emitted from the Sun contains sufficient energy when it reaches the Earth to
2 break (photolyze) chemical bonds in molecules, thereby leading to damaging effects on
3 living organisms and materials. Atmospheric O3 plays a crucial role in reducing exposure
4 to solar UV radiation at the Earth's surface. Ozone in the stratosphere is responsible for
5 the majority of this shielding effect, as approximately 90% of total atmospheric O3 is
6 located there over mid-latitudes. Ozone in the troposphere provides supplemental
7 shielding of radiation in the wavelength band from 280-315 nm, referred to as UV-B
8 radiation. UV-B radiation has important effects on human health and ecosystems, and is
9 associated with materials damage.
10 Adverse human health effects associated with solar UV-B radiation exposure include
11 erythema, skin cancer, ocular damage, and immune system suppression. A potential
12 human health benefit of increased UV-B exposure involves the UV-induced production
13 of vitamin D which may help reduce the risk of metabolic bone disease, type I diabetes,
14 mellitus, and rheumatoid arthritis, and may provide beneficial immunomodulatory effects
15 on multiple sclerosis, insulin-dependent diabetes mellitus, and rheumatoid arthritis.
16 Adverse ecosystem and materials damage effects associated with solar UV-B radiation
17 exposure include terrestrial and aquatic ecosystem impacts, alteration of biogeochemical
18 cycles, and degradation of man-made materials. Terrestrial ecosystem effects from
19 increased UV-B radiation include reduced plant productivity and plant cover, changes in
20 biodiversity, susceptibility to infection, and increases in natural UV protective responses.
21 In general, however, these effects are small for moderate UV-B increases at mid-
22 latitudes. Aquatic ecosystem effects from increased UV-B radiation include sensitivity in
23 growth, immune response, and behavioral patterns of aquatic organisms and the potential
24 for increased catalysis and mobility of trace metals. Biogeochemical cycles, particularly
25 the carbon cycle, can also be influenced by increased UV-B radiation with effects ranging
26 from UV-induced increases in CO2 uptake through soil respiration to UV-induced release
27 of CO2 through photodegradation of above-ground plant litter. Changes in solar UV
28 radiation may also have effects on carbon cycling and CO2 uptake in the oceans as well
29 as release of dissolved organic matter from sediment and algae. Finally, materials damage
30 from increased UV-B radiation includes UV-induced photodegradation of wood and
31 plastic.
32 There is a lack of published studies that critically examine the incremental health or
33 welfare effects (adverse or beneficial) attributable specifically to changes in UV-B
34 exposure resulting from perturbations in tropospheric O3 concentrations. While the
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1 effects are expected to be small, they cannot yet be critically assessed within reasonable
2 uncertainty.
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