Early evidence that governmental responses to COVID-19 reduce urban air pollution

There is no doubt that the global spread of COVID-19 represents the defining crisis of the last decade. Governments around the world have scrambled to try to reduce person-to-person spread and deal with pressures on public health infrastructure. Regions with community spread have almost universally faced restrictions on travel, business and social activities. These restrictions are designed to reduce the exponential spread of COVID-19 (that is, to flatten the curve), these restrictions will also have a large number of other economic, social and environmental repercussions. Here, I ask a simple question: Has reductions in economic activity and movement caused by governmental responses to COVID-19 improved air quality in cities? I compare February 2019 and 2020 air quality measures and show that six cities that were impacted early by government restrictions in response to COVID-19 show consistent declines in five of six major air pollutants compared to cities that were impacted later (the text in this post has been modified from Cadotte 2020).

One of the most pernicious and inevitable consequences of urbanization and industrialization is the release of air pollutants. The WorldHealth Organization (WHO) estimates that about 90% of urban residents experience air pollution that exceeds WHO guidelines and that air pollution is responsible for more than four million premature deaths annually (World Health Organization 2018). Air quality is adversely affected by the aerosol release of a number of chemical compounds from agriculture, manufacturing, combustion engines and garbage incineration, and is usually assessed by measuring the atmospheric concentrations of six key pollutants: fine particulate matter (PM_2.5), course particulate matter (PM₁₀), ground-level ozone (O₃), nitrogen dioxide (NO₂), sulfur dioxide (SO₂), and carbon monoxide (CO). These pollutants have a number of serious human health impacts (Table 1). Reducing inputs of these pollutants into urban areas requires a combination of technological advancement and behaviour change that can be stimulated by governmental regulations and incentives.

Table 1: The six commonly measured air pollutants in cities and their human health impacts.

Alterations of human, transport and industrial activity are usually the result of long-term economic and behavioural change and difficult to legislate under normal situations. However, the recent emergence of the global COVID-19 pandemic has had clear epidemiological impacts with, as of March 25, 2020, almost half a million confirmed infections and close to 20,000 deaths (World Health Organization 2020). This pandemic has resulted in emergency measures attempting to reduce transmission rates that limit activity, movement and commerce in jurisdictions around the world. While these emergency measures are critically important to limit the spread and impact of the coronavirus, they also provide a glimpse into how governmental calls for behavioural change can alter air pollution levels in cities.

Early evidence reveals that pollution levels have dropped in places that have undergone COVID-19 shutdowns. As Marshall Burke showed in a blog post,  PM_2.5 and PM₁₀, levels are lower than expected in parts of China. Here I examine January and February 2020 AQI levels for the six pollutants in Wuhan to what would be expected under normal circumstances. I further compare the change in February air pollution levels over the past two years in six cities that instituted emergency measures by the end of February (early impacted cities) to 11 cities that did not declare states of emergency until March (later impacted cities) using freely available air monitoring data (World Air Quality Index Project 2020) -see Table 2 for a list of cities.

Table 2: The eleven cities used in this analysis, the month that emergency measures were enacted and two- to six-year AQI averages of the pollutants

City-data come from monitoring agencies listed at the end of this post

Wuhan, China was the epicenter for the December 2019 emergence and the first person-to-person spread of the novel coronavirus.  In response, authorities initiated a series drastic measures limiting human movement and activity in Wuhan and large parts of Hubei province by the end of January. Three air pollutants: PM_2.5, PM₁₀ and NO₂ all showed substantial January and February declines in Air Quality Index (AQI) (U.S.Environmental Protection Agency 2014) values over 2019 levels for those months and what would be expected from long-term trends (Fig. 1). These long-term declining air pollution trends do reveal that China’s recentpollution reduction and mitigation efforts are steadily paying off, but the government-enforced restrictions further reduced pollution levels. The expected air pollution values predicted by temporal trends (red dashed lines in Fig. 1) are all substantially higher than the observed levels, with observed values being between 13.85% lower than expected for January PM_2.5 and 33.93% lower for January NO₂. Further, the reductions in the pollutants shown in Fig. 1 increased the number of days where pollutant concentrations were categorized as ‘good’ (0 < AQI < 50) or ‘moderate’ (51 < AQI < 100) according to the AQI. The three other pollutants: SO₂, O₃ and CO, all showed idiosyncratic or non-significant changes, mostly because their levels have already reduced significantly over time or appear quite variable (Fig. 2). 

Fig. 1. Temporal patterns of Air Quality Index (AQI) PM_2.5, PM₁₀ and NO₂ values in Wuhan, China. Both January and February, 2020 values show significant declines compared to 2019 levels and to that predicted from long-term trends (red dashed line).

Fig. 2. Temporal patterns of Air Quality Index (AQI) SO₂, O₃ and CO values in Wuhan, China.

Once COVID-19 moved to other jurisdictions, and confirmations of community spread emerged in February 2020, emergency measures, like those in Hubei province, were instituted to limit human movement and interaction. The cities subjected to February restrictions include, in addition to Wuhan, Hong Kong, Kyoto, Milan, Seoul and Shanghai, and the AQI values from these cities were compared to other cities that did not see the impacts of the novel coronavirus or have emergency restrictions in place until well into March. Log-response ratios between the air concentrations of pollutants observed in February 2020 to those from February 2019 reveal that all air pollutants except O₃ show a decline in the 2020 values for the early impacted cities (Fig. 3). For later impacted cities, there is no overall trend in changes in the concentrations of pollutants between 2020 and 2019 and the individual cities in this group showed less consistency in the differences between years (Fig. 3). 

Fig. 3. Log response ratios for Air Quality Index (AQI) PM_2.5, PM₁₀, NO₂, O₃, SO₂ and CO values between February 2019 and February 2020 values. Negative values indicate a decline in 2020. The green symbols indicate values from an assortment of cities that did not have emergency measures in place until March, 2020 (later impacted cities) and orange symbols are for cities that were impacted by the end of February.

These results indicate consistent air pollution reduction in cities impacted early by the spread of the novel coronavirus. However, the analyses presented here require further investigation as governments increasingly restrict activity world-wide, and some are discussing the possibility of prematurely lifting restrictions in order to spur economic growth. Further, the data analyzed here present point estimates of air quality but air pollution impacts are not homogeneous through urban landscapes and is influenced by spatial variation in industrial activities and transportation (Adams & Kanaroglou 2016). Thus, as higher resolution spatial air pollution data become available, it would be valuable to see how reduced activity affects air quality in different parts of cities.

This analysis of early data indicates that governmental policies that directly reduce human activity, commercial demand and transportation can effectively and quickly reduce urban air pollution. While the COVID-19 pandemic represents a serious risk for health and wellbeing of populations globally, especially those living in high density urban areas, the impacts of air pollution are equally consequential. If governments are willing to expend trillions of dollars in direct funding and indirect economic costs to combat this disease, then why do these same governments permit or even subsidize activities that emit air pollution? Maybe the lessons learned with COVID-19 can serve as the impetus for further action. Perhaps mandating changes to economic or transportation activity or investing in clean technology would better protect human health from the effects of air pollution.

Cited sources

Adams, M.D. & Kanaroglou, P.S. (2016) Mapping real-time air pollution health risk for environmental management: Combining mobile and stationary air pollution monitoring with neural network models. Journal of environmental management, 168, 133-141.
Cadotte, M. W. (2020) Early evidence that COVID-19 government policies reduce urban air pollution. Retrieved from eartharxiv.org/nhgj3

Cesaroni, G., Forastiere, F., Stafoggia, M., Andersen, Z.J., Badaloni, C., Beelen, R., Caracciolo, B., de Faire, U., Erbel, R. & Eriksen, K.T. (2014) Long term exposure to ambient air pollution and incidence of acute coronary events: prospective cohort study and meta-analysis in 11 European cohorts from the ESCAPE Project. Bmj, 348, f7412.

Fann, N., Lamson, A.D., Anenberg, S.C., Wesson, K., Risley, D. &Hubbell, B.J. (2012) Estimating the National Public Health Burden Associated with Exposure to Ambient PM2.5 and Ozone. Risk Analysis, 32, 81-95.

Greenberg, N., Carel, R.S., Derazne, E., Bibi, H., Shpriz, M., Tzur, D. & Portnov, B.A. (2016) Different effects of long-term exposures to SO2 and NO2 air pollutants on asthma severity in young adults. Journal of Toxicology and Environmental Health, Part A, 79, 342-351.

Kampa, M., & E. Castanas. (2008) Human health effects of air pollution. Environmental Pollution, 151, 362-367.

Khaniabadi, Y.O., Goudarzi, G., Daryanoosh, S.M., Borgini, A., Tittarelli, A. & De Marco, A. (2017) Exposure to PM 10, NO 2, and O 3 and impacts on human health. Environmental science and pollution research, 24, 2781-2789.

Raaschou-Nielsen, O., Andersen, Z.J., Beelen, R., Samoli, E., Stafoggia, M., Weinmayr, G., Hoffmann, B., Fischer, P., Nieuwenhuijsen, M.J. & Brunekreef, B. (2013) Air pollution and lung cancer incidence in 17 European cohorts: prospective analyses from the European Study of Cohorts for Air Pollution Effects (ESCAPE). The lancet oncology, 14, 813-822.

U.S. Environmental Protection Agency (2014) AQI: Air Quality Index. Office of Air Quality Planning and Standards, Research Triangle Park, NC.

World Air Quality Index Project (2020) https://waqi.info/.

World Health Organization (2018) Ambient (outdoor) air pollution: https://www.who.int/news-room/fact-sheets/detail/ambient-(outdoor)-air-quality-and-health.

World Health Organization (2020) Coronavirus disease 2019 (COVID-19), Situation Report –65.

City air quality monitoring agencies:

¹ Division of Air Quality Data, Air Quality and Noise Management Bureau, Pollution Control Department, Thailand (http://aqmthai.com).

² Delhi Pollution Control Committee (http://www.dpccairdata.com).

³ Hong Kong Environmental Protection Department (http://www.epd.gov.hk).

⁴BMKG | Badan Meteorologi, Klimatologi dan Geofisika (http://www.bmkg.go.id).

⁵South African Air Quality Information System - SAAQIS (http://saaqis.environment.gov.za).

⁶ Japan Atmospheric Environmental Regional Observation System (http://soramame.taiki.go.jp/).

⁷ UK-AIR, air quality information resource - Defra, UK (http://uk-air.defra.gov.uk).

⁸ South Coast Air Quality Management District (AQMD) (http://www.aqmd.gov/).

⁹ INECC - Instituto Nacional de Ecología y Cambio Climático (http://sinaica.inecc.gob.mx).

¹⁰ Agenzia Regionale per la Protezione dell'Ambiente della Lombardia (http://ita.arpalombardia.it).

¹¹ CETESB - Companhia Ambiental do Estado de São Paulo (http://cetesb.sp.gov.br).

¹² Department of Public Health of the Sarajevo Canton (http://mpz.ks.gov.ba/).

¹³ Air Korea Environment Corporation (http://www.airkorea.or.kr).

¹⁴ Shanghai Environment Monitoring Center (http://sthj.sh.gov.cn).

¹⁵ Israel Ministry of Environmental Protection (http://www.svivaaqm.net).

¹⁶ Air Quality Ontario - the Ontario Ministry of the Environment and Climate Change (http://www.airqualityontario.com/).

¹⁷ Wuhan Environmental Protection Bureau (http://www.whepb.gov.cn/).

Life in Plastic Ain’t so Fantastic

Guest post by Louis Vassos, MEnvSci Candidate in the Professional Masters of Environmental Science program at the University of Toronto-Scarborough

Much like the Buggles’ 1980 debut album, our material preferences are well within the age of plastic. Thanks to its light weight, durability, inertness, and low manufacturing costs, our use of plastics has increased dramatically since the mid-20^th century. From bottles and toys to car parts and electronics, there is seemingly no application beyond its reach. Despite its uses and benefits, it has come under increasing scrutiny by environmentalists in recent years. In this regard, we tend to think of larger-scale and more visible environmental impacts, such as accumulation in landfills and petrochemical use in manufacturing. There has also been a significant amount of research on plastic in marine environments, usually focused on larger debris known as macroplastics. Over the past decade, however, there has been increasing concern about a new type of plastic debris in our oceans. Though its presence was first highlighted in the 1970s, we are only just beginning to realize the impact of fragments known as microplastics. As their name would suggest, they are small pieces of plastic, typically measuring less than 5mm in diameter and sorted into two distinct classifications.

Primary microplastics are manufactured to be microscopically sized and are typically used in air blasting as a paint and rust remover, as well as in personal care products as an exfoliating scrubber. This latter use has risen sharply in cosmetics and facial cleansers since the 1980s, with plastic “microbeads” replacing natural materials such as pumice and ground almonds. Regardless of application they usually enter water bodies through drainage systems, and are easily able to pass through filtration systems at sewage treatment plants due to their small size.

Microbeads in toothpaste. Retrieved from: https://blog.nationalgeographic.org/2016/04/04/pesky-plastic-the-true-harm-of-microplastics-in-the-oceans/

Secondary microplastics arise from the breakdown of larger pieces of plastic debris on both land and in water. Larger debris will typically enter marine ecosystems directly or indirectly through careless waste disposal, often being transported through river systems. Sources of transfer include coastal tourism, extreme weather events, fishing, other marine industries, and accidental spillage during transportation. Over time, a culmination of processes such as exposure to UV radiation can reduce the debris’ structural integrity, causing brittleness, cracking, and yellowing. This in turn can lead to fragmentation through abrasion and waves, and fragments will gradually become smaller over time before reaching microplastic size (Cole et al, 2011).

As Eriksen et al (2014) have estimated, there is a minimum of 5.25 trillion plastic particles weighing 268,940 tons in the world’s oceans. Microplastics account for 92.4% of this mass, and their reach has been substantial. Because of their buoyancy and durability, they have the ability to travel long distances without degrading for years. Denser plastics (such as PVC) will sink and have the potential to reach coastal sediment (Andray, 2011). Other marine microplastics will end up trapped in ocean current systems known as gyres, the most famous grouping of which is the “Great Pacific Garbage Patch” in the North Pacific Gyre. Despite what the name would suggest, it is not an island-like mass of floating debris, but is more akin to an extensive “soup” of debris difficult to see with the naked eye. At a density of 334,271 pieces/km², microplastic mass in the area was found to be 6 times that of plankton (Moore et al, 2001). 

Potential microplastic transport pathways (From Wright et al, 2013)

Densities such as this increase potential microplastic ingestion by various marine organisms, especially filter feeders, plankton, and suspension feeders. These species may mistake debris for prey based on size or colour, or passively ingest them without being selective (Wright et al, 2013). In Farrell and Nelson’s (2013) study of mussel-eating crabs, they found that it is possible for microplastics to be transferred to individuals at a higher trophic level. The large surface area to volume ratio of microplastics makes them susceptible to water-borne pollutant contamination, and can cause toxic plastic additives such as BPA and PCB to leach into the water. This debris can also act as a dispersal vector for microbial communities, including potentially pathogenic species (Jiang et al, 2018). While the ingested debris can accumulate within individuals and be transferred up the food chain, the exact effects of this are not entirely known at this point in time (Avio et al, 2017). A recent study by Lei et al (2018), however, found that microplastics can cause oxidative stress and intestinal damage in zebrafish and nematodes, and that their toxicity is closely dependent on particle size.

Intestinal damage in zebrafish caused by exposure to 1.0 mg L^-1 of different microplastic types and sizes. Photograph A shows control (top), survival (middle), and dead after exposure (bottom) zebrafish (From Jiang et al, 2018)

Fluorescent microspheres on a crab’s gill lamella transferred from ingesting mussels, each measuring 5 micrometres in diameter (From Farrell and Nelson, 2013)

          What does the future hold for microplastics? Because their effects on both marine life and humans is relatively unknown, it is important to try and prevent them from entering and accumulating within marine environments. Properly dispose of larger plastic items to prevent them from entering waterways and breaking down into secondary microplastics, and be conscious about the presence of primary microplastics in other products. Make informed decisions when buying cosmetics, and choose ones that use natural exfoliating materials. Microbead bans have already begun to be enacted in several countries, including the UK, US, Canada and New Zealand (Pfeifer, 2018). There is also the potential for future studies on topics such as the health effects of microplastic ingestion and leached additives, debris behavior within the water column, and new standardized techniques for detection and sampling (Cole et al, 2011). It is hard to say what will happen next, but the removal of these 5.25 trillion particles from our oceans will prove to be a very difficult challenge without the development of novel extraction methods.

SOURCES

Anadrady, A.L. 2011. Microplastics in the marine environment. Marine Pollution Bulletin 62:1596 – 1605

Avio, C.G., S. Gorbi, and F. Regoli. 2017. Plastics and microplastics in oceans: from emerging pollutants to emerged threat. Environmental Research 128: 2 – 11

Cole, M., P. Lindeque, C. Halsband, and T.S. Galloway. 2011. Microplastics as contaminants in the marine environment: a review. Marine Pollution Bulletin 62:2588 – 2597

Eriksen, M., L.C.M. Lebreton, H.S. Carson, M. Thiel, C.J. Moore, J.C. Borerro. F. Galgani, P.G. Ryan, and J. Reisser. 2014. Plastic pollution in the world’s oceans: more than 5 trillion plastic pieces weighing over 250,000 tons afloat at sea. PLOS One

Farrell, P., and K. Nelson. 2013. Trophic level transfer of microplastic: Mytilus edulis (L.) to Carcinus maenas (L.). Environmental Pollution 177:1 – 3

Jiang, P., S. Zhao, L. Zhu, and L. Daoji. 2018. Microplastic-associated bacterial assemblages in the intertidal zone of the Yangtze Estuary. Science of the Total Environment 624:48 – 54

Lei, L., S. Wu, S. Lu, M. Liu, Y. Song, Z. Fu, H Shi, K. Raley-Susman, and D. He. 2018. Microplastic particles cause intestinal damage and other adverse effects in zebrafish Danio rerio and nematode Caenorhabditis elegans. Science of the Total Environment 619:1 – 8

Moore, C.J., S.L. Moore, M.K. Leecaster, and S.B. Weisberg. 2001. A comparison of plastic and plankton in the North Pacific Central Gyre. Marine Pollution Bulletin 42:1297 – 1300

Pfeifer, H. 2018. The UK now has one of the world’s toughest microbead bans. CNN. Retrieved from: https://www.cnn.com/2018/01/09/health/microbead-ban-uk-intl/index.html

Wright, S.L., R.C. Thompson, and T.S. Galloway. 2013. The physical impacts of microplastics on marine organisms: a review. Environmental Pollution 178:483 – 492

#EcoSummit2016: Conferences –the piñata of ideas.

One of the greatest benefits of attending conferences is that they represent learning opportunities. I don’t necessarily mean learning about new techniques or analyses, though you can undoubtedly find out about these at conferences, but rather conferences are opportunities to hear about new concepts, ideas and paradigms. In some ways conferences are like a piñata of ideas –they are chalk full of new ideas but you never know which you’ll pick up.

Ecosummit is not the typical conference I go to, it is much more diverse in topics of talks and disciplines of the attendees. This diversity –from policy makers, to social scientists, to ecologists, means that I am exposed to a plethora of new concepts. Here are a few nuggets that got me thinking:

• Knowledge-values-rules decision making context. Policy decisions are made at the interface of scientific knowledge, human values (what is important to people –e.g., jobs), and rules (e.g., economic laws). This seems like a nice context to think about policy, though it is not clear about how we prioritize new knowledge or alter values.

 

• Adaptation services. I work on ecosystem services (e.g., carbon storage, pollination support, water filtration, etc.), but I learned that ecosystems also provide adaptation services. These are aspects of ecosystems that will help human societies adapt to climate change (e.g., new products).

• Trees and air pollution. The naive assumption most of us make about trees in urban areas are that they improve local air quality. However, I saw a couple of talks where this may not necessarily be the case. Some species in North American (red oak, sweet gum, etc.) release volatile organic compounds. Spruce plantations may not take up nitrogen oxides, and in fact might release it. Thus we need to be careful on how we sell the benefits of urban trees.

• Transformative. This is a term I have certainly heard and used before, but in listening to a wide variety of talks, I realize it is used in different contexts to mean different things. I think it best to avoid this term.

• a-disciplinary.  I heard a guy say in a talk that he was a-disciplinary and so was not bound to the dogmas and paradigms of any discipline (I already have a hard time wrapping my head around interdisciplinary, multidisciplinary, transdisciplinary, etc.). He then presented a new paradigm and a number of prescribed well-formulated tools used to move from idea, communication, to action. I think the irony was lost on him.

The evolutionary canary in the coal mine*

*note -this post originally appeared on the Applied Ecologist's blog

Like canaries in coal mines, species can provide important information about deteriorating environmental conditions. A whole sub-discipline of environmental biomonitoring has emerged to provide the necessary tools to evaluate biological responses to changes in environmental conditions. While historically biomonitoring focused on contaminant concentrations in sentinel species –such as heavy metals in clams; modern biomonitoring uses information across multiple biological levels of organisation, from tissues, to organism behaviour, to the abundances and distributions of species. Since it is impossible to assess every aspect of an ecosystem’s response to pollution, scientists and practitioners still need to make decisions about which elements of an ecosystem should be monitored.

A coal miner with a canary –the classic sentinel species (url for photo: http://www.academia.dk/Blog/wp-content/uploads/CanaryInACoalMine_2.jpg)

In freshwater systems, diatoms are often the preferred organisms for monitoring since they have high diversity and diatom communities are structured strongly by local environmental conditions. Because of their long use in biomonitoring, freshwater biologists have sensitivity and indicator values for thousands of diatom species. Thus, in principle, you should be able to sample diatom communities in lakes and rivers of interest, and then assess the water quality based on the presence and abundance of different diatom species. While such proxies should always be validated and interpreted carefully (Stephens et al. 2015), there is a long and successful history of using diatoms for environmental monitoring.

Image of diatoms from a scanning electron microscope. (By Kostas Tsobanoglou - Own work, CC BY-SA 4.0, https://commons.wikimedia.org/w/index.php?curid=45315566)

The difficulty in practice is to identify diatom species, which requires expert training and can be time consuming. A number of researchers have pursued proxies and surrogates, for example using life form (e.g., diatom shape) or higher taxonomic groupings, instead of identifying species (Wunsam, Cattaneo & Bourassa 2002). In a recent article in the Journal of Applied Ecology, Francois Keck and colleagues (Keck et al. 2016) take this one step further, by using diatom evolutionary relationships as the biomonitoring tool.

Keck et al. employ novel statistical methods to create clusters of species based on their evolutionary relatedness from a phylogenetic tree and species’ sensitivity to pollution and show that these clusters, when delineated by short to moderate phylogenetic distances, do a good job of replicating species-level community pollution sensitivity indices.

This may seem like a onerous task, to assign diatoms to a correct position on a phylogenetic tree, but with the availability and now widespread use of DNA barcoding techniques, it is becoming easier to get genetic data for microscopic assemblages than to identify cells to species. This means that samples can be fit to the phylogenetic clusters without needing to shift through samples. Further, if species are observed, which have not been properly assessed for their sensitivity, they can be assigned an expected sensitivity value based on their relatedness to assessed species.

The phylogenetic tree and species’ sensitivities (Fig. 2 in Keck et al.).

While diatom evolutionary history may not have been strongly influenced by environmental pollutants in the past –because they are relatively recent stressors; it is clear from Keck et al.’s results that closely related species are similarly sensitive to pollution. Other fields of applied management have also begun to incorporate evolutionary history in the design and assessment of applied actions –for example, restoration (Hipp et al. 2015). Evolutionary history can provide important insights and management tools for dealing with the consequences of environmental change.

References

Hipp, A.L., Larkin, D.J., Barak, R.S., Bowles, M.L., Cadotte, M.W., Jacobi, S.K., Lonsdorf, E., Scharenbroch, B.C., Williams, E. & Weiher, E. (2015) Phylogeny in the Service of Ecological Restoration. American Journal of Botany, 102, 647-648.

Keck, F., Bouchez, A., Franc, A. & Rimet, F. (2016) Linking phylogenetic similarity and pollution sensitivity to develop ecological assessment methods: a test with river diatoms (microalgae). Journal of Applied Ecology.

Stephens, P.A., Pettorelli, N., Barlow, J., Whittingham, M.J. & Cadotte, M.W. (2015) Management by proxy? The use of indices in applied ecology. Journal of Applied Ecology, 52, 1-6.

Wunsam, S., Cattaneo, A. & Bourassa, N. (2002) Comparing diatom species, genera and size in biomonitoring: a case study from streams in the Laurentians (Quebec, Canada). Freshwater Biology, 47, 325-340.