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. 2020 Aug 25;47(17):e2020GL088913. doi: 10.1029/2020GL088913

Limited Regional Aerosol and Cloud Microphysical Changes Despite Unprecedented Decline in Nitrogen Oxide Pollution During the February 2020 COVID‐19 Shutdown in China

Michael S Diamond 1,✉, Robert Wood 1
PMCID: PMC7461095  PMID: 32904938

Abstract

Following the emergence of a novel coronavirus in Wuhan, China instituted shutdown measures starting in late January and continuing into February 2020 to arrest the spread of disease. This resulted in a sharp economic contraction unparalleled in recent Chinese history. Satellite retrievals show that nitrogen dioxide pollution declined by an unprecedented amount (~50% regionally) from its expected unperturbed value, but regional‐scale column aerosol loadings and cloud microphysical properties were not detectably affected. The disparate impact is tied to differential economic impacts of the shutdown, in which transportation, a disproportionate source of nitrogen oxide emissions, underwent drastic declines (~90% reductions in passenger traffic), whereas industry and power generation, responsible for >90% of particulate emissions, were relatively less affected (~20% reductions in electricity and thermal power generation). A combination of anomalously warm and humid meteorological conditions and complex chemical interactions further decreased nitrogen dioxide concentrations but likely enhanced secondary aerosol formation.

Keywords: COVID‐19, aerosol, pollution, cloud, china, emissions

Key Points

  • Rapid economic collapse in February 2020 due to the COVID‐19 pandemic and partial recovery in March 2020 are unique in recent Chinese history

  • Regional decline in NO2 pollution is unprecedented in the satellite record, but no change in aerosol or cloud properties is detected

  • Different economic impacts by sector, meteorology, and complex chemical interactions help explain differences in NOx and aerosol response

1. Introduction

1.1. Emergence of a Novel Coronavirus and the Societal Response to the Resulting COVID‐19 Pandemic

In late December 2019, cases of a pneumonia of unknown cause were reported in the city of Wuhan. By January 2020, the pathogen responsible—a novel zoonotic coronavirus—had already spread throughout China and other nations including Japan, South Korea, and the United States (C. Wang, et al., 2020). To arrest the spread of COVID‐19 (the disease caused by the novel coronavirus), a series of strict restrictions on travel and other activities were adapted across China, slowing the spread of the epidemic in China even as the disease became a global pandemic (Maier & Brockmann, 2020; Tian et al., 2020).

Unsurprisingly, this socioeconomic “shutdown” had a catastrophic effect on the Chinese economy. Figure 1 shows the purchasing managers' index (PMI) for both manufacturing and nonmanufacturing sectors as reported by the National Bureau of Statistics of China. The PMI is a survey‐based estimate of economic activity, with values above 50% corresponding to growth and below to contraction (Harris, 1991). February 2020 stands out sharply, featuring a decline in manufacturing PMI deeper than any point during the aftermath of the 2008 financial crisis and the only period of contraction in nonmanufacturing PMI since records for that index began in 2007, followed by a rapid recovery.

Figure 1.

Figure 1

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Economic and environmental indicators from January 2005 to May 2020. (a) China's purchasing managers' index. Economic growth is indicated by blue shading and contraction by red shading (darker colors signify the manufacturing and lighter colors the nonmanufacturing sectors). (b) Monthly anomalies of ln(NO2) (red) and aerosol optical depth (blue) averaged over eastern China and cloud droplet effective radius (green) averaged over the East China Sea. February 2020 is highlighted in gray shading. (c) Year‐on‐year growth rate of passenger transportation by month. (d) Combined January–February totals of power generation and some heavy industries.

1.2. Anthropogenic Drivers of Recent Pollution Changes

Nitrogen oxides (NOx ≡ NO + NO2) are reactive, short‐lived gases that are a major constituent of air pollution harmful to human health (Atkinson et al., 2018). Due to rapid cycling between NO and NO2 in the atmosphere, changes in NO2 indicate changes in NOx overall. Tiny particles suspended in the air (aerosol) are another major component of air pollution. Particulate matter with aerodynamic diameters of 2.5 μm or smaller (PM2.5) is known to have severe health impacts, with some estimates of global excess mortality due to outdoor PM2.5 approaching 10 million deaths annually (Burnett et al., 2018).

In addition to their relevance for public health, aerosol particles influence the climate through absorbing and scattering sunlight and by changing the optical properties of other components of the Earth system like snow, ice, and clouds. In particular, gaps in our knowledge about the interactions between clouds and aerosol particles represent the largest source of uncertainty in present‐day anthropogenic radiative forcing in the most recent Intergovernmental Panel on Climate Change (IPCC) assessment (Myhre et al., 2013). Aerosol particles can serve as cloud condensation nuclei (CCN) upon which liquid‐phase cloud droplets may form. Increasing CCN increases the number of cloud droplets and (for the same amount of liquid water) decreases their size, increasing cloud reflectivity (Twomey, 1977). Macrophysical cloud adjustments to these microphysical changes can either enhance or offset this “Twomey effect” (Ackerman et al., 2004; Albrecht, 1989). Aerosol also influences mixed‐phase and ice cloud properties, although the climatic effects of these changes are even less certain (Storelvmo, 2017).

Determining whether observed cloud changes are attributable to aerosol or meteorological factors is a major challenge (Gryspeerdt et al., 2016; Stevens & Feingold, 2009). To better constrain causality, there has been a growing literature on “natural experiments” like volcanic eruptions and inadvertent anthropogenic modifications like ship tracks (Malavelle et al., 2017; Toll et al., 2019). A clear signal in aerosol and cloud properties due to the February 2020 shutdown would be of great interest to those working to constrain the magnitude of aerosol‐cloud interactions.

Both long‐term changes in air pollution and aerosol‐cloud interactions over China and short‐term changes attributable to individual events have been analyzed extensively. Increasing aerosol and cloud droplet number concentrations were associated with China's rapid economic growth between the 1980s and the 2000s (Bennartz et al., 2011). Over the past decade, however, sulfate aerosol and cloud droplet number concentrations have declined (McCoy et al., 2018). Recent declines in pollutants like PM2.5 and NO2 have been driven by increasingly stringent environmental policies (Jin et al., 2016) including goals set by the eleventh (2006–2010) and twelfth (2011–2015) Five‐Year Plans (de Foy et al., 2016; Liu et al., 2016; van der A et al., 2017), the 2013 Air Pollution Prevention and Control Action Plan (A. Ding et al., 2019; Silver et al., 2018; Zhai et al., 2019; Zhang et al., 2019; B. Zheng et al., 2018; Y. Zheng et al., 2017), and the 2018 Three‐Year Action Plan for Winning the Blue Sky Defense Battle. Ephemeral pollution decreases have also been associated with high‐profile events like the 2008 Olympics and Paralympics in Beijing (Cermak & Knutti, 2009; Witte et al., 2009), the 2010 World Expo in Shanghai (Hao et al., 2011), and the 2014 Youth Olympic Games in Nanjing (J. Ding et al., 2015).

Research has already begun on the environmental consequences of COVID‐19. Carbon dioxide emissions have significantly declined due to shutdown measures worldwide (Le Quéré et al., 2020). Strong declines in NO2 have been observed in Europe, China, South Korea, and the United States (Bauwens et al., 2020; Liu et al., 2020); however, emissions reductions related to the early Chinese shutdown were insufficient to avoid bad haze episodes in several cities (Chang et al., 2020; X. Huang et al., 2020; Le et al., 2020; P. Wang et al., 2020) and benefits from decreasing NOx and PM2.5 may be offset by related increases in ozone (X. Huang, Ding, Gao, et al., 2020; Shi & Brasseur, 2020). We add to this work by examining regional‐scale NOx changes alongside possible aerosol and cloud changes in the context of the 2005–2020 satellite record and by drawing upon meteorological, economic, and emissions data to help explain the differences we observe between pollutants during China's February 2020 shutdown.

2. Data

We analyze monthly mean pollution and cloud properties from January 2005 to May 2020 using National Aeronautics and Space Administration's (NASA's) “A‐train” satellite constellation (local overpass times ~13:30). NO2 and SO2 data are from the Ozone Monitoring Instrument (OMI) on Aura (Levelt et al., 2018; Schoeberl et al., 2006) and aerosol optical depth (AOD), liquid‐phase cloud droplet effective radius (r e), single‐scatter albedo (ω 0), and Ångström exponent (α) data are from the Moderate Resolution Imaging Spectroradiometer (MODIS) on Aqua (Parkinson, 2003). For OMI, we compute monthly averages using NASA's standard daily 0.25° by 0.25° gridded products for NO2 tropospheric column retrievals screened for cloud fractions below 30% (Krotkov et al., 2017, 2019) and planetary boundary layer (PBL) SO2 (Krotkov et al., 2015). For MODIS, we use standard monthly 1.0° by 1.0° gridded Collection 6.1 products for AOD at 550 nm, liquid r e retrieved using the near‐infrared 2.1 μm channel and a visible channel (Hubanks et al., 2019; Platnick et al., 2017), and ω 0 at 470 nm and α from the Deep Blue algorithm (Hsu et al., 2013; Sayer et al., 2019). More information about the satellite retrievals are provided in the supporting information.

Two‐meter air temperature and specific humidity and 10‐m winds are from the Modern‐Era Retrospective analysis for Research and Applications, Version 2 (MERRA‐2) product (Gelaro et al., 2017) and are bilinearly interpolated to OMI/MODIS grids as needed.

Monthly economic statistics are compiled from the National Bureau of Statistics of China (NBSC). Emissions of NOx, PM2.5, and SO2 in 2015 broken down by economic sector (IPCC, 2006) are provided by the Emissions Database for Global Atmospheric Research (EDGAR), version 5.0 (Crippa et al., 2018, 2020).

3. Pollution Changes During the February 2020 Shutdown

3.1. Regression Model

Monthly anomalies of ln(NO2), AOD, and r e are shown in Figure 1b. We average the ln(NO2) and AOD values over a region (20–42°N, 108–125°E) encompassing the eastern provinces of the People's Republic of China, Hong Kong, and Taiwan. We average r e over a region (23–35°N, 122–130°E) in the East China Sea in which previous studies have examined aerosol‐related cloud microphysical trends (Bennartz et al., 2011; McCoy et al., 2018). Perhaps surprisingly, no clear perturbation can be seen in the cloud or aerosol fields in February 2020, and it is not clear that the apparent decline in ln(NO2) is statistically distinguishable from other variations not associated with such an abrupt socioeconomic upheaval. Simply comparing observed pollution levels in 2020 to values from previous years or a long‐term average can be misleading due to seasonal cycles and trends in emissions and/or concentrations. For example, one may expect February 2020 pollution levels to have been below the 2015–2019 mean even without any coronavirus‐related disturbance due to policy‐related trends.

To address these issues, we employ an ordinary least squares linear regression model (Pedregosa et al., 2011) to better estimate what monthly mean air pollution and cloud properties should have been in the absence of coronavirus‐related perturbations:

Ytϕλ=c0ϕλ+cpreϕλPpret+cpostϕλPpostt+cCNYϕλHCNYt+ccosϕλScost+csinϕλSsint, (1)

where Y is the geophysical variable of interest—ln(NO2), AOD, or r e—as a function of time (t, in months since January 2005) estimated independently at each grid box of latitude ϕ and longitude λ and the regressors P, H, and S are defined as follows. Each constant c and the intercept c 0 are calculated independently at each grid box. We use the natural logarithm of NO2 rather than its absolute value for the regression because NO2 is distributed log‐normally in space (de Foy et al., 2016) and later analyses involve taking spatial averages. The model is fit based on data from January 2005 to December 2019 and is used to predict the expected values for January–May 2020.

We define the regressors P pre and P post as the trend in terms of months preceding or following January 2013, respectively. January 2013 was chosen as the policy “turning point” due to the severe haze events that occurred early that year (R. Huang et al., 2014; G. Zheng et al., 2015) that contributed to the creation of the ambitious Air Pollution Prevention and Control Action Plan in September 2013 (Jin et al., 2016).

Next, the regressor H CNY refers to whether the Chinese New Year holiday and related festivities (Jiang et al., 2015; Tan et al., 2009) occurred during time t. When Chinese Lunar New Year begins in February, H CNY is assigned a value of 1. When the holiday begins in late January with festivities lasting into early February, H CNY is assigned a value based on the fraction of the 10 days following the Lunar New Year occurring in each month. H CNY is set to zero all other times.

Finally, the regressors S cos and S sin refer to idealized seasonal cycles, represented as two annual Fourier modes:

Scost=cos2πmt−112, (2)
Ssint=sin2πmt−112, (3)

where m refers to the calendar month (January = 1) at time t.

Supporting information Figure S1 shows the value of each regressor as a function of time; Figure S2 shows maps of the coefficient of determination (R 2), root‐mean‐square (RMS) error (ε RMS), and number of samples used to fit the regression (N); and Figures S3–S5 show maps of the intercept and regression coefficients for ln(NO2), AOD, and r e, respectively.

3.2. Results

Maps of the observed (retrieved) and expected (regression) values of NO2, AOD, and r e for February 2020, along with the difference between the observed and expected values, are shown in Figure 2. (NO2 values are converted into their absolute values for the sake of presentation.) There is a readily apparent decline in NO2, in some places rivaling the magnitude of the total tropospheric column, consistent with the results of Bauwens et al. (2020). However, there are no consistent differences in either the aerosol loading or cloud microphysics on a regional scale. Furthermore, regions with the largest discrepancies in the cloud fields tend to be those with the lowest explained variance (Figure S2). The lack of a regional AOD decline is in apparent conflict with some previous results (X. Huang, Ding, Gao, et al., 2020; Le et al., 2020; Shi & Brasseur, 2020), although it should be noted that the quantities of interest (regional‐scale column aerosol loading versus surface PM2.5 at specific stations) are different. There does appear to be an increase in AOD surrounding Beijing, which is consistent with some previous results (X. Huang, Ding, Gao, et al., 2020; Le et al., 2020). The only area with what appears to be a coherent decline in AOD occurs around Shanghai, although even that city experienced relatively high levels of surface PM2.5 in two events earlier in the month (Chang et al., 2020).

Figure 2.

Figure 2

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Change in pollution over China for February 2020. (left column) Observed values, (middle column) estimated values, and (right column) their difference are shown for (top row) NO2, (middle row) AOD, and (bottom row) r e. Shading is such that lighter (left and middle columns) and greener (right column) colors align with what would be expected for less pollution. Areas without valid data are shaded in gray. Gray stippling indicates absolute differences below 2ε RMS. Black boxes in the right column indicate areas used for the regional averages.

To look at the differences between the observed and estimated values in greater historical perspective, we average the observed and estimated ln(NO2), AOD, and r e values over the same regions used in Figure 1b. Results are displayed in Figure 3, with the root‐mean‐square (RMS) error in this case calculated using the differences between the spatially averaged observed and expected values from January 2005 to December 2019. Like the PMI indices, ln(NO2) shows a pronounced and unprecedented decline in February 2020 followed by a rapid recovery. In contrast, AOD and r e values during 2020 are not perceptibly distinct from the 2005–2019 record.

Figure 3.

Figure 3

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Time series of observed (Obs) and expected (Exp) values and their differences for (a, b) ln(NO2), (c, d) AOD, and (e, f) r e, as averaged over the boxes in Figure 2. Manufacturing PMI is shown for reference.

Sulfate aerosols, which can form via the oxidation of SO2 in the atmosphere, are a major source of cloud condensation nuclei. PBL SO2 as retrieved by OMI is noisier than the tropospheric column NO2 retrievals and can be affected by factors like elevated volcanic plumes, although previous analyses have successfully analyzed regional‐scale trends to gain insight into long‐term pollution changes (Krotkov et al., 2016; McCoy et al., 2018). Figure S6 shows PBL SO2 averaged over eastern Asia excluding major Northern Hemisphere volcanic eruptions. SO2 levels in February 2020 are in line with the decreasing trend since 2013 and are slightly above the previous year's values, consistent with the minimal change in cloud and aerosol properties.

3.3. Detectability

The differences between our model's expectations and the observations can be due to emissions changes not captured in our broad policy‐related trend terms or other factors such as meteorological variability. Thus, for a given emissions reduction, we may not be able to reliably detect that reduction given other fluctuations.

We estimate the probability of detection, defined as the probability that we would observe a regionally averaged change exceeding a given threshold, as follows. For emissions reductions up to 90%, we adjust the expected value by a proportional amount for NO2 and AOD and then generate a spread using a normal distribution fit with the adjusted expectation and RMS error (Figures 3b, 3d, and 3f). The fraction laying below the threshold is the probability of detection. This method assumes that concentrations/optical properties decline proportionally to emissions, which neglects the idea that not all emissions are anthropogenic; that there may be complex interactions between emissions and concentrations; and that optical properties may not vary linearly with aerosol concentration (especially if the mix of sources changes) but should still provide useful results assuming that natural emissions are small compared to the anthropogenic emissions and if interpreted cautiously. A similar method is employed for r e, taking into account the more indirect relationship between aerosol and cloud properties by estimating the increase in droplet size for a given reduction in aerosol number (N a) as

δlnre≈δlnNaδlnNdδlnNaδlnreδlnNd≈−βδlnNa3, (4)

where N d is the cloud droplet number concentration, β is the log‐log sensitivity of N d to aerosol changes, and the factor of −1/3 relating r e and N d changes assumes liquid water path changes are negligible. Because the β parameter is highly uncertain, we bound plausible r e responses by making separate estimates for low and high sensitivities of 0.3 and 0.8, respectively (Bellouin et al., 2020).

Figure S7 shows the probability of detection for ln(NO2), AOD, and both bounded estimates for r e for detection thresholds of decreases (NO2 and AOD) or increases (r e) of one, two, and three RMS errors of the (unadjusted) expected value for February 2020. If emissions were reduced by greater than 80%, there is a very large probability that we would have detected changes in all three variables at the strictest (3ε RMS) threshold (unless the clouds have very low susceptibility to aerosol), whereas if emissions were reduced by 20% or less, we would be unlikely to detect changes in any of the variables at the strictest threshold and only have even odds of detecting aerosol or cloud changes at the most lenient threshold (1ε RMS). For a 50% emissions reduction, we would be very likely to detect a NO2 change at any threshold, whereas detection of changes in AOD or r e (assuming higher sensitivity) would only be very likely at the 1–2ε RMS thresholds and have even odds at the 3ε RMS level. Given that the AOD and r e perturbations we estimated are both within 1ε RMS of their expected values, this analysis suggests that aerosol reductions during the February 2020 lockdown period were unlikely to have exceeded ~20%.

4. Factors Influencing Pollution Changes During the February 2020 Shutdown

4.1. Meteorology

Meteorology is an important driver of changes in pollution concentrations (de Foy et al., 2016; P. Wang et al., 2020; Zhai et al., 2019; Zhang et al., 2019). To assess the effects of meteorology on our results, we create a second linear regression model as follows:

Ymettϕλ=c0ϕλ+cpreϕλPpret+cpostϕλPpostt+cCNYϕλHCNYt+cTϕλT2Mtϕλ+cqvϕλQ2Mtϕλ+cuϕλU10Mtϕλ+cvϕλV10Mtϕλ, (5)

where Y met is the geophysical variable of interest, the terms relating to policy‐driven trends and holiday effects are the same as for the model in Equation 1, and the idealized seasonal cycle has been replaced by the meteorological terms T 2M for 2‐m air temperature, Q 2M for 2‐m specific humidity, and U 10M and V 10M for 10‐m zonal and meridional winds, respectively. All meteorological variables are standardized by subtracting the 2005–2019 mean from each monthly value and dividing by the 2005–2019 standard deviation. Figure S8 shows maps of regression diagnostics (R 2, ε RMS, and N), and Figures S9–S11 show maps of the intercept and regression coefficients for ln(NO2), AOD, and r e, respectively.

Figure S12 shows time series of the regionally averaged results, which are overall similar to those in Figure 3.

The differences between the model described by Equation 1 and the observations can be interpreted as being due to a combination of meteorology and emissions (including emissions‐concentrations‐chemistry interactions), whereas the difference between the model described by Equation 5 and the observations can be interpreted as being due to emissions alone. (Of course, there also may be other factors that have not been accounted for or meteorological effects not captured by the linear model.) The effect of meteorology can be estimated in two ways: as the difference between the expectations between the two models or as the difference between the model described by Equation 5 and the same model with the meteorological factors set to their mean values for a given month. Similarly, the (linear) contribution of individual meteorological variables can be estimated as the difference between the model described by Equation 5 and that model with the meteorological variable of interest set to its monthly mean value.

Figure S13 shows the decomposition of emission and meteorology effects using this framework for February 2020. Meteorology alone would have led to a ~10% decrease in NO2, driven mostly by temperature, followed by humidity. The emissions effect alone would have led to a ~50% decrease. In contrast, meteorology is estimated to have increased AOD by ~5%. Meteorological anomalies during February 2020 are shown in Figure S14. Conditions were warmer and wetter than average, corresponding with increased chemical sinks of NO2 but more potential for secondary production of aerosol (Le et al., 2020).

4.2. Emissions

Another explanation for the different NOx and aerosol responses is that economic sectors that disproportionately emit one or the other pollutant may have been impacted differently by the shutdown. Figures 1c and 1d show changes in passenger transportation, energy generation, and iron and steel production from 2005 to 2020. Figure S15 includes a variety of other economic subsectors tracked by the National Bureau of Statistics of China. Passenger transportation, in particular, was devastated by the shutdown. In contrast, total January–February power generation was down ~10% (similar to 2008–2009), implying a decrease of ~20% in February alone. Heavy industries like steel production (slightly up) were comparatively unaffected, with the remainder of the economy somewhere in between.

Different reporting metrics were chosen based on data quality and availability. For a more directly comparable (but temporally limited) perspective, Figure S16 shows the change in January–March production between 2020 and 2019 for each subsector.

Anthropogenic emissions of NOx ( ENOx), PM2.5 ( EPM2.5), and SO2 ( ESO2) for the year 2015 from EDGAR are combined to create aggregate “transportation” and “industry and power” sectors, with the remainder lumped into an “other” category primarily consisting of agriculture and waste management. (See supporting information Tables S1–S3 for the specific breakdown of the transportation, industry and power, and “other” sectors, respectively, by IPCC 2006, code.) Figure S17 shows maps of the contributions of the three pollutants by economic sector. Transportation is a major source of NOx pollution, comparable to the industry and power sectors, whereas the industry and power sectors dominate emissions of primary PM2.5 and SO2 (a major aerosol precursor). Supporting information Tables S1–S3 provide sums of annual emissions in 2015 over eastern China for each sector grouping. The transportation sector accounts for 26.2% of all NOx emissions but only 4.7% of PM2.5 and 3.6% of SO2 emissions (Table S1), while the industry and power sectors account for 72.3% of NOx, 92.8% of PM2.5, and 95.1% of SO2 emissions (Table S2). Given that decreases in NOx, PM2.5, and SO2 since 2015 have been driven primarily by regulations targeting the industrial sector (Zhang et al., 2019), it is likely that the transportation sector comprised a greater share of total NOx emissions in 2020 than in 2015.

The ~50% decline in NO2 due to emissions can be mostly explained by a dramatic fall in transportation emissions paired with a 10–20% reduction in energy and industrial sources. A ~20% decline in aerosol emissions would likely go undetected by our method (Figure S7). At the same time, this analysis suggests that the estimated ~40% decline in coal burning during the Chinese lockdown period from Le Quéré et al. (2020) is unrealistically high, both because electrical and thermal power declined by only ~20% at peak and because a ~40% decline in aerosol emissions should have been detectable, at least at the less stringent thresholds. The Chinese coal data in that study come from a private industry analysis of several power plants, which may not have been representative of the national response. However, given that reductions in ground transportation dominate their analysis as well, even halving the industrial change estimates would not substantially change their results (e.g., the peak global decline on 7 April 2020 would fall ~10%, from ~17 to ~15 MtCO2/day).

One caveat about our aerosol estimates is that we assume that the significant aerosol reductions between 2013 and 2017 continued during 2018–2020. If we instead were to freeze P post at its 2017 value (assume no further progress), the observed AOD would have been 10–20% below its expected value. Our analysis is consistent with no aerosol change, a moderate decrease, or, as discussed below, a moderate decrease in emissions compensated by increased secondary production.

4.3. Chemistry

As a further complication, meteorological anomalies, emissions changes, and their interactions influenced atmospheric chemistry and therefore pollution concentrations in February 2020.

During the winter, the atmospheric lifetime of NOx (~1 day) over eastern China decreases with decreasing emissions as higher ozone concentrations allow for more loss via reaction with hydrogen oxide radicals (HOx) during the day and via hydrolysis of N2O5 (an important NOx reservoir) within aerosols at night (Shah et al., 2020). We may therefore expect decreases in NO2 concentrations to exceed reductions in NOx emissions.

PM2.5 concentrations may either increase or decrease as a response to a NOx decline depending on its magnitude and background concentrations (Zhao et al., 2017). The increase in O3 as NOx emissions fall (Shi & Brasseur, 2020) increases HOx concentrations and thus the atmospheric oxidizing capacity, which could further reduce NO2 but facilitate secondary aerosol formation (X. Huang, Ding, Gao, et al., 2020). This in combination with the relatively warm and wet February 2020 meteorological conditions offers a compelling explanation for the apparent increase in aerosol surrounding Beijing (X. Huang, Ding, Gao, et al., 2020; Le et al., 2020; P. Wang et al., 2020), although the effect appears to be weaker in other regions, perhaps due to differences in background conditions. However, regional aerosol changes are not necessarily independent as transport and aerosol‐boundary layer interactions can lead to a complex interplay in pollution levels between regions (Chang et al., 2020; X. Huang et al., 2020).

To assess potential changes in aerosol composition and size distribution, we analyze ω 0 and α using the model described by Equation 1. Figure S18 shows time series of observed and expected values, and Figure S19 shows maps of the February 2020 anomalies. Since 2013 there has been a trend toward higher ω 0 (less absorption) and lower α (coarser size distribution), which is reversed during February 2020 (more absorbing and finer aerosol). The anomalies in February 2020 are similar to those in December 2019, however, so it is not clear that these anomalies can be attributed to shutdown effects.

5. Conclusions

Despite unprecedented declines in economic activity and NOx emissions during the February 2020 coronavirus shutdown in China, we find no detectable perturbation in aerosol and related cloud properties. The severe curtailment of passenger transportation (a disproportionate NOx source) but comparatively muted changes in power generation and heavy industry (disproportionate PM2.5 and SO2 sources), along with meteorology and complex chemical interactions, help explain this discrepancy.

Further study of the environmental consequences of COVID‐19 is warranted, not least because potential links between long‐term and short‐term air quality and vulnerability to the disease remain unresolved (Contini & Costabile, 2020). There is some evidence that short‐term exposure to air pollution increased the case fatality rate of the 2002–2003 Severe acute respiratory syndrome outbreak in several Chinese cities (Cui et al., 2003), which raises the possibility of feedbacks between containment measures that reduce pollution and population‐level resilience. Additionally, dramatically reduced transportation sector emissions without similar changes in other sectors could represent a plausible future emissions mix if widespread electrification of transportation is adopted but other sectors do not adopt similar pollution mitigation measures.

Supporting information

Supporting Information S1

Click here for additional data file. (19MB, pdf)

Acknowledgments

M. S. D. was supported by NASA Headquarters under the NASA Earth and Space Science Fellowship Program Grant NNX‐80NSSC17K0404. We thank Kyle Armour, Fiona Lo, Daniel McCoy, Joel Thornton, and Casey Wall for helpful discussion and Aijun Ding and Jean‐François Müller for constructive reviews that helped improve the manuscript.

Diamond, M. S. , Wood, R. (2020). Limited regional aerosol and cloud microphysical changes despite unprecedented decline in nitrogen oxide pollution during the February 2020 COVID‐19 shutdown in China. Geophysical Research Letters, 47, e2020GL088913 10.1029/2020GL088913

Data Availability Statement

OMI/Aura and MODIS/Aqua Level 3 gridded data and MERRA‐2 meteorological reanalysis data are publicly available from NASA's Goddard Earth Sciences Data and Information Services Center (https://disc.gsfc.nasa.gov/). Various economic statistics from the People's Republic of China are publicly available from the National Bureau of Statistics of China (http://www.stats.gov.cn/english/). EDGAR's annual sector‐specific grid maps are publicly available from the European Commission's Joint Research Center (https://edgar.jrc.ec.europa.eu/overview.php?v=50_AP; https://data.europa.eu/doi/10.2904/JRC_DATASET_EDGAR).

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Associated Data

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Supplementary Materials

Supporting Information S1

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Data Availability Statement

OMI/Aura and MODIS/Aqua Level 3 gridded data and MERRA‐2 meteorological reanalysis data are publicly available from NASA's Goddard Earth Sciences Data and Information Services Center (https://disc.gsfc.nasa.gov/). Various economic statistics from the People's Republic of China are publicly available from the National Bureau of Statistics of China (http://www.stats.gov.cn/english/). EDGAR's annual sector‐specific grid maps are publicly available from the European Commission's Joint Research Center (https://edgar.jrc.ec.europa.eu/overview.php?v=50_AP; https://data.europa.eu/doi/10.2904/JRC_DATASET_EDGAR).

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