Wildfire and prescribed burning impacts on air quality in the United States

Introduction

The 2020 Critical Review (CR) by Jaffe et al. (2020a) provides an in-depth examination of large-scale outdoor biomass burning locations and intensities throughout the United States. It documents remote and ground-based measurement technologies and data bases that are used to examine these fires both in near real-time and historically, observing an increasing annual trend in the areas burned. Adverse effects on human health, visibility, and climate are considered. Particulate matter (PM), carbon monoxide (CO), ozone (O₃) precursors, oxides of nitrogen (NO_x), and more than 500 organic compounds, many with potential for adverse health effects, are discussed. Fires are also large contributors to the Earth’s carbon dioxide (CO₂) load. The CR summary (Jaffe et al. 2020b) is supplemented with articles on wildfire and air quality management, data bases, and health effects (Baker et al. 2020; DeMerritt 2020; Freeburn 2020; Kinsman 2020; Lahm and Larkin 2020; Sasser and Lahm 2020) in the June issue of EM magazine. Herein, expert discussants provide additional perspectives and information on the topic. Their appearance as coauthors does not necessarily indicate their agreement with the opinions of other discussants. An online supplement contains additional information and illustrations related to this discussion.

Organic composition of biomass burning aerosols-Qi Zhang

Biomass burning (BB) emissions are enriched in hazardous air pollutants, CO, NO_x, PM, and many organic compounds (Andreae and Merlet 2001; Bond et al. 2004). Organic aerosols (OA) are major BB emission components. BB organic aerosols (BBOA) comprise directly emitted primary organic aerosols (POA) and secondary organic aerosols (SOA) produced from the oxidation of volatile organic compounds (VOCs) (Johansson et al. 2016; Koss et al. 2018). BBOA chemical compositions, physical properties, and toxicity vary with fuel types, burning conditions, and atmospheric aging.

BB POA derives from the thermal degradation of cellulose, hemicellulose, and lignin – plant-derived biopolymers which together account for nearly 90% of the fuel’s dry weight. Major pyrolysis by-products are levoglucosan and related anhydrosugars, such as mannosan and galactosan (Simoneit 1999). Levoglucosan is a chemical marker for BB emissions and its abundance indicates BB contributions to ambient OA as well as the extent of atmospheric aging as its abundance decreases (Cubison et al. 2011). Lignin is a class of biopolymers containing phenolic hydroxyl and methoxy functional groups; its thermal breakdown releases phenols such as guaiacol (2-methoxyphenol), syringol (2,6-dimethoxyphenol), and their derivatives (Nolte et al. 2001; Schauer et al. 2001; Simoneit 1999). Although most of the BB phenols are semi-volatile or intermediately-volatile, and mainly in the gas phase, these compounds can undergo fast reactions in both gas- and aqueous-phases to form SOA with high mass yields (Smith et al. 2014; Sun et al. 2010). Phenolic SOA products are diverse, including oligomers and associated derivatives, aromatic species containing various hydroxyl, carboxyl, carboxylate, carbonyl, and alkyl functional groups, and compounds derived through convoluted bond-breaking and bond-forming reactions (Yee et al. 2013; Yu et al. 2014, 2016). Phenols react with nitrogen dioxide, nitrous acid, and nitrate radicals – oxidants that are abundant in urban downwind areas due to intense NO_x emissions. These reactions produce nitrophenols which are of concern for their toxicity and light absorbing properties (Pang et al. 2019; Yuan et al. 2016; Zhang et al. 2016). BBOA also includes nitroaromatics, dioxins (e.g., polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans, and polychlorinated biphenyls), and polycyclic aromatic hydrocarbons (PAHs) (Fleming et al. 2020; Samburova et al. 2016; Zhang, Buekens, and Li 2017).

Flaming combustion converts fuel carbon to CO₂ at >90% efficiency and releases lower amounts of organic gases, particles, and CO compared to smoldering combustion (Akagi et al. 2013; Bertrand et al. 2018; Chen et al. 2007; Collier et al. 2016; McMeeking et al. 2009). BBOA in smoldering plumes is more complex than that from flaming, with the two phases distinguished by the modified combustion efficiency (MCE), determined by the molar ratio of CO₂ to the sum of CO₂ and CO. Detailed chemical characteristics of BBOA as a function of MCE for natural wildfires are still being determined.

Atmospheric aging changes BBOA composition and properties depending on reactant composition and meteorological conditions; a wide gap in knowledge remains concerning this aging. Laboratory burns show SOA-forming potential (Grieshop, Donahue, and Robinson 2009a, 2009b; Hennigan et al. 2011; Ortega et al. 2013). Yet little to no enhancement of BBOA relative to CO is seen in real-world burns, even though BBOA shows clear signs of aging and oxidation over time (Capes et al. 2008; Jolleys et al. 2012; Zhou et al. 2017). One hypothesis is that SOA formation is balanced by dilution and subsequent evaporation of semi-volatile organics during atmospheric aging (May et al. 2015). A more thorough understanding of the BBOA aging processes is necessary for improving climate models, developing air pollution control strategies, and improving understanding of human health implications.

Health effects of biomass burning-Michael T. Kleinman

Previous A&WMA critical reviews and discussions (Chow et al. 2006; Eklund et al. 2014; Ferris 1978; Goldstein 1983; Grahame, Klemm, and Schlesinger 2014; Mukerjee 1998; Pope and Dockery 2006; Vedal 1997; Watson et al. 1997) examined relationships between pollutant concentrations and health, but they are not specific to BB exposures. These exposures occur sporadically so that most general population health effects studies are retrospectively conducted with ecological time series analyses. Studies of acute cardiovascular and/or respiratory responses from smoke exposures are discussed in a review on firefighter exposure (Adetona et al. 2016). Most of the cited studies analyzed lagged effects and were restricted to within six days of exposure. Study outcomes include incidences of mortality, hospital admissions, physician or emergency room visits due to events or symptoms. Symptoms resulted from chronic obstructive pulmonary disease (COPD), asthma and cardiovascular episodes such as stroke, heart failure and cardiac dysrhythmia, i.e., outcomes that are mostly relevant to people with preexisting diseases. For cardiovascular diseases, many study findings were not significant with respect to hospital admissions, physician or emergency room visits (Crabbe 2012; Deflorio-Barker et al. 2019; Delfino et al. 2008; Duclos, Sanderson, and Lipsett 1990; Martin et al. 2013). However, smoke exposure events are associated with increased hospital admissions for respiratory conditions (Reid et al. 2016). Same day admissions for chronic obstructive lung disease (Martin et al. 2013) and asthma (Deflorio-Barker et al. 2019; Martin et al. 2013) were significantly associated with smoke exposure. Logistic regressions relating fire smoke and non-smoke PM_2.5 exposures with overall and age-stratified cardiovascular and respiratory diseases found statistically significant associations for asthma and combined respiratory disease, but no significant relationships for cardiovascular diseases (Stowell et al. 2019). Stowell et al. (2019) found positive age-specific associations with smoke exposure for asthma and combined respiratory disease in children, and for asthma, bronchitis, COPD, and combined respiratory disease in adults. No significant associations were found in older adults. A caution is that interactions between high temperatures and wildfire smoke exposure are associated with increased deaths from cardiovascular, respiratory, and nervous system diseases (Shaposhnikov et al. 2014). Since wildfire frequencies and severities are expected to increase due to climate change, the additive or synergistic effects of combined elevated temperature and smoke exposures should be considered in health risk assessments.

Kim et al. (2018) compared mutagenicity emission factors (EF in revertants/kg fuel) for a variety of fuel types during flaming and smoldering, finding that the smoldering EFs were factors of ten or more than those of the flaming phase. Although flaming smoke appeared to have higher revertants/µg of PM, the much higher emission rates from smoldering resulted in increased exposure.

Air quality tradeoffs between wildfires and prescribed burns-Fernando Garcia-Menendez

The CR notes that fire emission factors differ as a consequence of varying fire intensities, fuels, and combustion efficiencies. Beyond emissions, transport and transformations of smoke from wildfires and prescribed burns are also distinct, resulting in differences in health impacts associated with each type of fire. While wildfires expose large populations to severe air pollution episodically, prescribed fire smoke can impact local populations with lower concentrations over longer time intervals (Navarro et al. 2018). Fire provides a balance in many ecosystems, and past suppression efforts have upset this balance (North et al. 2015; Snider, Daugherty, and Wood 2006).

An obstacle to comparing wildfire and prescribed burn effects is that they cannot be evaluated in isolation, but must be considered as coupled components of an environmental system (Hunter and Robles 2020). Regular prescribed fire treatments under selected weather conditions can mitigate aggregate air pollution associated with wildland fire (Kochi et al. 2012; Schweizer, Cisneros, and Buhler 2019; Zhao et al. 2019). Air quality analyses that incorporate the consequences of prescribed fire on wildfire likelihood, burn area (i.e., fire leverage, Loehle 2004), intensity, and emissions are needed to understand the tradeoffs of wildfire and prescribed fire emissions. Williamson et al. (2016) define wildfire and prescribed fire “smoke regimes” and their associated public health impacts. Beyond the feedbacks between prescribed fire and wildfire related to fire dynamics, such analyses should consider the implications of plume transport, atmospheric chemistry, smoke exposure, and impacted populations. Piecing together the required components, spanning different research fields, and their interactions at broad scales may be one of the largest challenges for long-term smoke management.

Air quality impacts must also be placed into the larger context of land and environmental management. Decisions about fire suppression and fuel treatment not only influence fire damage and air pollution; they also determine land management costs and environmental benefits tied to wildland conditions (Kline 2004). Air quality is only one dimension of fire management decision-making in which public health and ecological land management goals can be misaligned. Comprehensive cost-benefit analyses of wildland fires that consider monetized fuel treatments, fire and smoke damages, ecological services, climate benefits, and other components of complete land management programs must be developed (Mercer, Haight, and Prestemon 2008). A more uniform land management approach across the nation is desirable, though challenging.

Visibility, fires, and regional haze-Tom Moore

The supplemental material provides background information on visibility metrics, emphasizing how light extinction is determined from the weighted sum of chemical composition and extinction efficiencies. Organic and elemental carbon (OC and EC) aerosols have high extinction efficiencies, so emissions from wildfires and prescribed burns have a large effect on regional haze. Figures discussed here can be accessed in the supplement.

Biomass burning sources are categorized (WRAP 2001) as wildfire (Wx), prescribed wildland fire (Rx), agricultural crop burning (Ag), and residential wood combustion (RWC, mostly for space heating). Open burning of refuse and structural fires are considered to be minor contributors. Biogenic emissions from agriculture, natural and human-controlled releases from the biosphere, and evergreen forests, are SOA precursors, but are not considered as part of smoke management. Wx are considered to be natural emissions, even when an individual fire results from a human cause. Rx, Ag, RWC, and the other minor sources are anthropogenic sources and considered to be controllable.

The Regional Haze Rule (U.S.EPA 2017) and its planning guidance (Tsirigotis 2019) intend to separate natural from anthropogenic contributions to light extinction to better track long-term progress on visibility improvement with emission reductions, especially those of state-regulated sulfur dioxide (SO₂) and NO_x precursors to the ammonium sulfate and ammonium nitrate concentrations, which Figure S-3 shows are important contributors throughout the U.S. Much of the western U.S. has cleaner air and better visibility on the “Most Impaired Days” in Figure S-3. High concentrations obviously affected by Wx fires are removed from these analyses, but less intense Wx fire contributions are still included. Atmospheric clarity combined with the distance at which an object can be discerned from the background are affected by even minor amounts of biomass burning smoke relative to smoke-free conditions. Combined OC and EC contributions in the Montana to New Mexico tier of states and westward, where more than 75% of the parks and wilderness areas protected under the Regional Haze program are located, constitute 30% to 50% or more of the observed impairment. Smoke contributions vary from year-to-year and in time and space. Increasing frequencies and intensities of unmanageable Wx fires documented in the CR appear to be offsetting progress from SO₂ and NO_x emission reductions (McClure and Jaffe 2018). Western U.S. Wx acres burned exceed those for Rx and Ag fires by a factor of ten.

Smoke Management Programs (SMPs) have been established for Rx and Ag fires by state, tribal, and local air quality agency regulatory programs in cooperation with federal or state land managers and/or other landowners. Examples of these programs are shown in Figure S-4 and Table S-1. The Clean Air Act does not require SMPs to address Rx and Ag fires and their regulation varies across the country. Western U.S. lands where Rx burns are applied are predominantly owned by federal land management agencies such as the U.S. Forest Service, Dept. of Interior Bureau of Land Management, National Park Service, and Fish and Wildlife Service, as well as other federal agencies, state public lands agencies, with some private ownership. Ag fires are concentrated in areas such as eastern WA, north-central ID, the Central Valley of CA, and other agricultural production areas around the West, on both private and tribal lands.

Rx and Ag acres burned and emissions vary by season, safety considerations, fuel loadings, weather, and potential effects on air quality. SMP structures and operations anticipate the number, timing, size and potential emissions of burn events. Burn permissions are requested from the SMP, and are approved or modified for the “next day” based on evaluation of the factors cited above and potential for interstate smoke transport into the area. The “actual” vs. “approved” emissions from an Ag or Rx burning event are estimated based on the number of acres and the amount of fuel consumed.

SMP operations do not systematically prevent or mitigate Wx fire activity or impacts. Many of the Wx fires are human-caused due to unintended ignitions, and the sizes and persistence of these fires are affected by climate change and accumulated fuel loadings due to land management decisions over the past century. SMP Wx fire mitigation efforts will require improved methods to optimize where and when Wx-prone areas require preventative measures.

Regulatory perspective-Meryln Hough

The following topics warrant further emphasis from a regulatory perspective: 1) Air Quality Index (AQI) reporting and advisories; 2) use of low-cost air quality sensors; 3) social media and website updates; 4) exceptional event guidance and limitations; and 5) community safe spaces and resiliency planning.

Wildfires provide opportunities to educate the public on the AQI, which is used throughout the U.S. to report and forecast air quality. The AQI embodies hazards from multiple pollutants, with an AQI of 100 equal to the National Ambient Air Quality Standard (NAAQS) for PM, O₃ or another criteria pollutant. Airshed strategies are designed to keep air quality in the Green – Good and Yellow – Moderate categories, but wildfires have pushed the AQI into the unhealthy to hazardous (Red, Purple, and Maroon) regions in the Pacific Northwest. The AQI has been helpful for communicating air quality health advice to athletic directors and event organizers. During the 2008 Olympic Trials for Track & Field in Eugene, OR, northern California wildfires contributed heavy smoke aloft to the area. Thunderstorms mixed the smoke to lower elevations that increased population exposure. After reviewing the AQI-based health advice with the air quality agency, Trials organizers postponed events when the AQI exceeded 150 (Red – Unhealthy category). September 2017 fire-related AQIs were worse, in the Purple – Very Unhealthy and even in the Maroon – Hazardous categories. The Oregon Health Authority translated the AQI into a brochure with specific advice for outdoor school activities. Football practices and games were postponed, relocated or canceled as a result of this advice.

Over 40 inexpensive light scattering PM sensors (PurpleAir 2020) were deployed in Lane County, OR, for less than the cost of a single compliance monitor; thereby expanding the geographic network and providing air quality information for site-specific event decisions. Default PM_2.5 calibrations for these sensors yielded concentrations exceeding those from collocated compliance monitors, and their outputs should be used in a relative rather than absolute sense.

Air quality agencies are updating and expanding website capabilities to keep the public and other stakeholders informed. EPA’s AirNow tool allows comparisons of local air quality with larger regions. Facebook and Twitter feeds are reposted by the news media and other agencies, thus keeping everyone better informed. Without the expanded website tools and social media options, agency staff can be overwhelmed with public demand for information. On September 5th, 2017, our agency received over 73,000 web hits – 300 times the normal amount. The Oregon smoke blog website (Table S-1) reports wildfire locations and nearby air quality levels. Morning updates involve federal, state and local forestry representatives, air agencies and health departments that coordinate public messaging and outreach. The CR recognizes the Blue Sky model and deployment of Air Resource Advisors in recent years which are advancing effective communication of wildfire impacts.

As noted in the CR, U.S.EPA (2016) Exceptional Events (EE) guidance intends to prevent penalizing areas for air quality situations outside their control. The concept of “regulatory significance” requires modification. Our agency submitted nine days during 2017 that measured PM_2.5 exceeding 35 µg/m³ during wildfires. Only two of the nine days were approved by EPA as EEs since that was sufficient to bring the 2015–2017 data into compliance with the 3-year standard. However, the 2015 and 2016 concentrations were low due to favorable meteorology. More typical weather and PM_2.5 levels during 2018 and 2019 made clear that all of the wildfire-affected 2017 concentrations were needed to attain the 2017–2019 NAAQS. The eliminated 7 EEs had retrospective regulatory significance. Excluding documented EEs removes a bias from air quality trends that would demonstrate the effectiveness of necessary emission reduction measures, such as controversial home wood heating control measures. Current EE guidance does not achieve the intended goal “to prevent penalizing communities for events outside their control.”

Urban airshed strategies have focused on ambient (or outdoor) air quality to meet NAAQS and Clean Air Act deadlines. The increasing frequency and severity of wildfire impacts have caused greater concerns about indoor air quality during these events. There is a need to identify and provide safe indoor spaces, and include these as part of the overall emergency resilience planning, as noted in the CR.

A 24-hour NAAQS masks higher diurnal concentrations- Eric Stevenson

The CR notes successes and difficulties with smoke forecasting and modeling. An additional dimension is the “democratization” of air monitoring with lower cost sensors that has made the public more aware of air quality, but with limited understanding of how air quality is measured and communicated. Continuous PM compliance monitors provide hourly and, in some instances sub-hourly measurements which are supplied to the public. Explaining the intricacies of averaging times is challenging and may complicate messaging that is needed for the public to take appropriate actions. The 2017 northern California wildfire smoke transported to the San Francisco Bay Area affected many events, including professional, collegiate and amateur sports. Athletes and spectators were concerned about high outdoor exposures during these events. Health officials also wanted a “bright line” on when actions should be taken. If forecasters predicted high concentrations, how wide an area should be notified regarding potential cancelation of events or modification of activities? What actions should schools and other facilities housing sensitive populations take, as buildings have different filtering capabilities to remove concentrations from indoor air?

AIRNOW (2020) has developed an algorithm called “NowCast” that uses previous hourly concentrations to predict concentrations at a given monitoring location for the coming 24-hours with access to the public for many locations. This is a blunt tool that does not include complex meteorology and other factors that may affect that 24-hour average concentration. In most circumstances, health departments are responsible for releasing “stay at home” or other action orders. Air quality professionals need to interact with health departments prior to these fire events to determine the best course of action for their areas. Air quality and health professionals also need to provide clear, consistent, informative, and actionable messages. Being proactive is key until methods are accepted that address sub-daily concentrations.

Indoor air, forest fuel loading reductions, and O₃ increases during wildfires- Samuel L. Altshuler

Sheltering-in-place is promoted to reduce exposure to outdoor air contaminants, but if the sheltering locations have indoor emissions, often the case within homes, exposures may also be high. In homes with combustion sources (e.g., gas stove tops, fire places, wood stoves, candles, etc.), there is a potential for poor air quality. High efficiency HEPA air filters, HVAC filters with ratings >13, and indoor air cleaners can mitigate indoor air pollution as well as smoke intrusions. Use of solvents or strong cleaners (containing bleach) indoors should also be avoided. Use of N-95 masks outdoors should also reduce PM exposures.

Biomass harvesting to reduce fuel loadings can be used as a renewable energy source. This would reduce both CO₂ and toxic emissions. Figure S-5 contrasts biofuel produced from undergrowth clearing in California’s Sierra Nevada ranges with a prescribed burn pile at Big Trees State Park, less than 2 km away. Greater coordination among nearby landowners could make better use of a valuable biomass resource.

Although PM_2.5 is the main component of fire emissions, effects on O₃ should not be neglected, as noted in the CR. Figures S-6 and S-7 show hourly concentrations of PM_2.5 and O₃ at air quality stations in the Bay Area during the 2017 California wildfires. Early in the event PM_2.5 concentrations were elevated, but not at their maxima. O₃ attained higher concentrations, particularly in the East Bay urban area of Hayward (0.14 ppm), than had been observed for over a decade. As PM_2.5 concentrations increased toward the end of the event, O₃ levels dropped below 0.10 ppm. The event also coincided with a record heat wave on September 1st and 2nd. This is consistent with the CR observation that O₃ impacts are highest at moderate PM_2.5 levels.

Just as different woods have different tendencies to create creosote in fireplace flues, wildfires of different biomass species produce differing organic emission, as noted above. Kim et al. (2018) found that the greatest lung toxicity was from eucalyptus, which is representative of chaparral-type wood. They also found that smoke from pine wood, which is broadly distributed across the United States, caused genetic mutations in bacteria, an indicator for cancer. Emissions from fires in regions rich in these fuels may induce greater health effects than those from fires of similar magnitude with other types of trees.

Peat burning, measuring aged smoke, and brown carbon for exceptional events-Judith C. Chow and John G. Watson

While forest fires are the main source of biomass smoke in North America, peat burning is a major global source, with the largest emissions from Southeast Asia (Indonesia and Malaysia). Although peatlands occupy only ~3% of the Earth’s land surface, they store 20–30% of the planet’s terrestrial carbon (Page, Rieley, and Banks 2011; Rein, Cohen, and Simeoni 2009). Once started, peat fires last for weeks to months, are nearly impossible to extinguish, and are dominated by the smoldering-phase (Hu et al. 2018; Stockwell et al. 2016), resulting in higher toxic emissions, as discussed above. Although the CR and this discussion have some relevance to peat burning, its emissions and management needs to be treated in another comprehensive review.

Recent peat-burning emissions tests (Chow et al. 2019a; Watson et al. 2019) illustrated how multipollutant emission aging from biomass burning and other sources can be simulated with an oxidation flow reactor (OFR) (Cao et al. 2020). The OFR is like a miniature smog chamber that irradiates an emissions mixture with high intensity ultraviolet radiation to simulate photochemical changes during pollution transport and aging. As noted by Dr. Zhang, aging does not necessarily increase PM_2.5 mass concentrations and does change chemical composition. Mass losses are evident in the decrease of the low-temperature carbon fraction (OC1 evolving at temperatures <140°C) (Chow et al. 2007), consistent with the suggested hypothesis of semi-volatile compound evaporation. The levoglucosan source marker abundances are greatly reduced, making it less useful for quantitative source apportionment.

The long-term U.S. IMPROVE and CSN PM_2.5 speciation networks initiated the multi-wavelength carbon analysis method in 2016 (Chow et al. 2015) which allows the brown carbon component of biomass smoldering to be determined (Chow et al. 2019b, 2018). Such information may be useful for determining the amount of non-manageable contributions from biomass burning during exceptional events, as explained by Mr. Hough.

Final remarks-Daniel A. Jaffe

The topic of air quality and fires is complex and multi-disciplinary. For this reason, no one expert can cover the entire range of important topics. The 2020 CR was fortunate to have assembled an outstanding team, including Susan O’Neill, Sim Larkin, Amara Holder, Dave Peterson, Jessica Halofsky, and Ana Rappold. A&WMA provided the authors a wide latitude to complete this CR. Each of these outstanding scientists applied their expertise and knowledge to the problem to generate what we hope is a useful product for the scientific and policy communities. The constructive feedback from the Critical Review committee and this interchange with the discussants adds value to the effort. I am grateful to the entire team as it has been a pleasure to work with this great group of scientists.

Supplemental data

Supplemental data for this paper can be accessed on the publisher’s website.

Additional information

Notes on contributors

Samuel L. Altshuler

Samuel L. Altshuler is an Air Quality Consultant and current Past Chair of the Critical Review Committee.

Qi Zhang

Qi Zhang is a Professor in the Department of Environmental Toxicology, at the University of California, Davis, CA

Fernando Garcia-Menendez

Fernando Garcia-Menendez is an Assistant Professor of Environmental Engineering at North Carolina State University

Charles Thomas (Tom) Moore

Charles Thomas (Tom) Moore is Air Quality Program Manager at the Western Regional Air Partnership/Western States Air Resources Council, Ft. Collins, CO

Merlyn L. Hough

Merlyn L. Hough is an environmental engineer and Executive Director of the Lane Regional Air Protection Agency in Eugene-Springfield, Oregon, and a member of the Critical Review Committee

Eric D. Stevenson

Eric D. Stevenson is former Director of Meteorology and Measurement, Bay Area Air Quality Management District, San Francisco, CA and incoming Chair of the Critical Review Committee

Judith C. Chow

Judith C. Chow is a Research Professor at the Desert Research Institute, past Chair of the Critical Review Committee, and author of the 1995 Critical Review

Daniel A. Jaffe

Daniel A. Jaffe is Professor of Atmospheric Chemistry at the University of Washington, Seattle, WA, and author of the 2020 Critical Review

John G. Watson

John G. Watson is a Research Professor at the Desert Research Institute, past Chair of the Critical Review Committee, and author of the 2002 Critical Review.