Reducing emissions from ongoing anthropogenic activities such as fossil fuel production and intensive ruminant farming is essentially palliative: it helps but does not cure. More substantial reductions can be achieved by ending or substantially reducing the activities themselves, for example, by shifting entirely away from coal and gas as energy sources and by reducing ruminant emissions. But such actions require political debate and major social adjustments, which are outside the scope of this journal. The more limited mitigation options discussed here are less contentious and can be implemented within the framework of existing commitments to the UNFCCC Paris Agreement.

U.K. national inventory of methane emissions by source sector. Note the significance of the decline in emissions from waste and energy sources in the late 1990s, but the relative lack of progress in recent years. From UK NAEI ().

The global challenge is clear from Table 1 . The need is to cut emissions from fossil fuels and from biomass and biofuel burning and reduce the methane footprint of agriculture and waste. But within this global envelope, each nation has its own spectrum of emissions. Figure 2 illustrates the impact of a sustained effort to reduce U.K. national emissions. Earlier emission cuts were very significant, but recently, the reduction trend has slowed or ceased. In 2017, the National Atmospheric Emissions Inventory (UK NAEI) estimated total U.K. emissions from waste as about 756 kilotons (kt) (quoted in tons by NAEI with no errors given: 1 kt = 0.001 Tg), fugitive emissions from energy fuels as 217 kt, 63.5 kt from combustion, and 187 kt from agriculture. If emissions from waste, fuels, and combustion were removed, the national total could be cut a further 80% or more.

A challenge to policy‐making is that there is a major discrepancy (Leip et al., 2018 ; Saunois et al., 2016 , 2019 ) between “top‐down” estimates (Nisbet & Weiss, 2010 ) of the annual global methane emission, assessed from measuring the atmosphere, and “bottom‐up” totals summing emissions estimates based on national data (e.g., number of cows and area of rice fields): the “bottom‐up” numbers are typically much higher (e.g. EDGAR‐ European Commission, Joint Research Centre / Netherlands Environmental Assessment Agency, 2011 ). Top‐down results can cause major revisions to previous emission inventories, for example after the work of Bergamaschi et al. ( 2010 ). Press reports of an earlier study by this team show that it caused substantial revisions ( https://www.theguardian.com/environment/2006/jun/22/climatechange.climatechangeenvironment ). Chang et al. ( 2019 ) also illustrate this problem: they found, compared to previous estimates, a larger than expected contribution of ruminant enteric emissions to the increasing trend in global methane emissions between 2000 and 2012. This discrepancy may point to errors in either or both top‐down and bottom‐up assessment methodologies, a problem that does not instill confidence in using bottom‐up emissions estimates to identify the optimum (least cost, most benefit) targets for reduction efforts.

For example, Canada has committed to reduce methane emissions from the oil and gas sector by 40–45% by 2025 relative to 2012 levels (Canada, 2016 ). To take another example, in the United Kingdom where landfill emissions have been vigorously reduced, the national atmospheric emissions inventory suggests methane emissions have dropped by 61% since 1990 (UK NAEI, 2019 ; UK NIR, 2019 ), a pattern seen in many developed nations. In these specific sectors, there has been much progress in abatement of emissions, at relatively low cost by improved industrial practices to detect and prevent leaks, or to capture and use methane before it is emitted to atmosphere. But for many sources with more diffuse emissions that are less easily detected (as local air has low incremental methane over background), there has been little progress.

Many jurisdictions have brought in legislation to control methane emissions (Iacobuta et al., 2018 ). In 2019, the Net Zero report of the U.K.'s Committee on Climate Change led to a national commitment to reduce net national greenhouse emissions to zero by 2050 (UK CCC, 2019 ). New Zealand has passed a similar promise. Though these commitments focus on CO 2 , most nations also intend to reduce methane emissions in addition to cutting CO 2 emissions. The focus has been on point sources where very high mole fractions of methane can be measured in the nearby air—such as gas leaks, or gas emissions around landfills. Such sources can typically be detected easily, precisely located and then mitigated.

Globally, the rate of growth in fossil fuel methane emissions may be slowing (Schaefer et al., 2016 ). Nisbet et al. ( 2016 , 2019 ) used isotopic evidence to infer that fossil fuel emissions have declined as a proportion of the total methane budget, though they could not rule out an increase in absolute terms. But it is also possible (within the uncertainties of the observations) that a decline in biomass burning may be masking a rise in fossil fuel emissions (Worden et al., 2017 ).

Cutting methane emissions has long been identified as a low‐hanging fruit for climate action (Hansen et al., 2000 ; UNEP, 2018 ). Shindell et al. ( 2017 ) considered as much as 110 Tg CH 4 per year of abatement was possible by scaling up existing technology and industrial best practice and policy options, to provide societal benefits that outweigh implementation cost. A significant cut in anthropogenic methane emissions would be sufficient to reduce the overall methane burden within decades or less, to thus return to the RCP2.6 pathway. There is wide benefit in reducing methane emissions (Boucher & Folberth, 2010 ) including major reduction opportunities at less than zero cost (Warner et al., 2015 ).

(a) A national inventory report of anthropogenic emissions by sources and removals by sinks of greenhouse gases, prepared using good practice methodologies accepted by the Intergovernmental Panel on Climate Change and agreed upon by the Conference of the Parties serving as the meeting of the Parties to this Agreement; and

“Parties, noting the importance of technology for the implementation of mitigation and adaptation actions under this Agreement and recognizing existing technology deployment and dissemination efforts, shall strengthen cooperative action on technology development and transfer.” And that “Accelerating, encouraging and enabling innovation is critical for an effective, long‐term global response to climate change and promoting economic growth and sustainable development.”

The amount of methane in the atmosphere above preindustrial levels directly affects the allowable cumulative carbon emissions budget (i.e., the total anthropogenic CO 2 emitted before reaching a temperature threshold, usually 1.5 or 2 °C). The rise thus has major implications for the UN Paris Agreement's goal to limit climate warming to 2 °C and especially to the 2018 Intergovernmental Panel on Climate Change (IPCC) advice to limit warming to 1.5 °C (IPCC, 2018 ). The global warming consequences of this rise, relative to the Paris‐compliant path, are shown in Figure 1 from Nisbet et al. ( 2019 ). Moreover, the unexpected growth in methane has already significantly negated the expected impact of progress in controlling CO 2 emissions (Nisbet et al., 2019 ). Haustein et al. ( 2017 ) suggest that methane is contributing to the accelerating global warming rate at a time when CO 2 emissions may be stabilizing. If growth in the methane burden continues at current rates in the coming decades, it will become impossible to meet the Paris target, especially in view of the recent upward reevaluation of the global warming potential of methane (Etminan et al., 2016 ).

Methane and the Paris Agreement. Top panel shows evolution of the mean global atmospheric mixing ratio (green open circles) compared to Representative Concentration Pathway RCP2.6 (Collins et al.,) as used in the last IPCC Assessment and consistent with the Paris Agreement. RCP2.6 concentrations peaked in 2012, whereas those for the RCP 4.5 (dashed) pathway peak around 2045, while RCP 8.5 (dotted) mole fractions continue to rise throughout the 21st century. Lower panel shows differences in radiative forcing from the RCP2.6 pathway and actual evolution of methane (green open circles), CO(red circles), and nitrous oxide (blue circles). Dashed lines show RCP 4.5 relative to RCP 2.6. Radiative forcing used here is from Etminan et al. (), and the figure is updated from that in Nisbet et al. () using NOAA mole fraction data to June 2019.

One illustrative pathway that would lead to compliance with the 2015 United Nations (UN) Paris Agreement of the United Nations Framework Convention on Climate Change (UNFCCC, 2015 ) is Representative Concentration Pathway 2.6 (RCP 2.6) (Collins et al., 2013 ; Rogelj et al., 2012 ) (Figure 1 ). This pathway envisaged an immediate and significant fall in methane, allowing time for progress on the more difficult task of reducing CO 2 . This is not a unique pathway to achievement of the Paris goals, but being Paris compliant is thus commonly cited as an exemplar of the reduction pathways needed. Yet, by 2019, as a result of the unexpected rise, the atmospheric methane mixing ratio was over 100 ppb above the RCP 2.6 path (Nisbet et al., 2019 ). If added to the atmosphere as a single pulse, this would equate to an input of about 300 Tg of methane (1 ppb is globally equivalent to about 2.8 Tg CH 4 ).

The rise in atmospheric methane has been accompanied by a significant shift in the C‐isotope ratio of atmospheric methane to values more depleted in 13 C (Nisbet et al., 2016 , 2019 ). The post‐2007 growth and isotopic shift has now been sustained for over a decade, but the reasons for the change are still not well understood. Possible explanations for this growth and the concurrent isotopic shift (for isotopic values of sources, see Sherwood et al., 2017 ); may include increases in biogenic emissions (especially in the tropics and subtropics), changes in the chemical sinks of methane by atmospheric OH or Cl, increased fossil fuel emissions coupled with declining biomass burning (Nisbet et al., 2019 , 2016 ; Schaefer et al., 2016 ; Schwietzke, Sherwood et al., 2017 ; Turner et al., 2017 , 2019 ; Rigby et al., 2017 ; Worden et al., 2017 ), more oxidation of methane by methanotrophy in forest soils (Ni & Groffman, 2018 ), or, more likely, some combination of all these factors.

Since 2007 the atmospheric methane burden has risen sharply. It began growing by about 6 ppb/yr in 2007, after a sustained pause at the start of the millennium. The growth rate accelerated in 2014 and is currently around 10 ppb/yr (Nisbet et al., 2014 ; Nisbet et al., 2016 ; Nisbet et al., 2019 ). Note that the words “concentration” and “level,” terms derived from wet chemistry, are often used in popular media when discussing methane. The gas‐specific terms “mole fraction” and “mixing ratio” are preferred for use in this report.

Methane is a remarkably attractive target for reducing atmospheric greenhouse warming. It is the secondmost important anthropogenic greenhouse gas, and a major contributor to the increase in tropospheric ozone, also a major greenhouse gas. Methane has a 100‐yr global warming potential of 32 compared to CO 2 (Etminan et al., 2016 ); thus, removing 1 Mg (a ton) of methane has real impact (Collins et al., 2018 ). Moreover, with an average tropospheric lifetime of about 9–10 yr (Dlugokencky et al., 2011 ), successful moves to reduce methane emissions can have rapid effectiveness.

A much larger event followed 3 yr after Elgin, when a blowout took place from the Aliso Canyon underground storage facility in California in 2015. This emission was studied by a team which included participants both from the effort to quantify the Elgin leak and also from the study of fracking in U.S. gasfields (Conley et al., 2016 ). In an impressive campaign, they flew dozens of plume transects over the region, as well as taking whole air samples on the ground downwind. This successfully mapped out the dispersion of the leaked gas. However, in contrast to the Elgin study, which used δ 13 C CH4 in an attempt to identify the source reservoir, Conley et al. ( 2016 ) measured ethane, C 2 H 6 , as the marker to distinguish the emissions from the natural gas leak from other sources of methane.

Isotopic measurements (Figure 9 ) then helped pinpoint the source of the gas in the blowout. Keeling plots in air samples from flights transecting the gas plume (Figure 9 ) found the source gas had δ 13 C CH4 of ‐42.3 ± 0.7‰. The results pointed to a source in a small gas pocket, geologically shallower than the main gasfield and that the potential volume of gas involved in the leak was limited. The hypothesis was borne out. The leak diminished rapidly. This FAAM aircraft study informed the platform operators, advising them that the gas flux from leak was reduced. The leak was then ended by accessing the platform for a top kill, rather than drilling down for a much slower bottom kill that had been discussed.

Case Study B: unexpected events: Elgin gas leak . Aircraft are important when major unexpected events occur. An uncontrolled gas release took place from the Elgin platform in the U.K. North Sea from 25 March to 16 May 2012 (Lee, Mobbs, et al., 2018 ). When the gas blowout occurred, the U.K. FAAM (Facility for Airborne Atmospheric Measurement) BAe146 aircraft was fortuitously fully equipped as it had just returned from measuring methane in Scandinavia: It was immediately sent to fly over the gas platform (Figure 9 ). Flying at low altitude, FAAM was able to map out and sample the plume of methane‐rich gas from the leak (Lee, Mobbs, et al., 2018 ). This study was carried out flying normal to wind direction. As a general rule however, where the source is known, it is better to fly at an angle (e.g., 30°) to wind direction in order to maximize the intersection.

(left) Elgin gas production platform, from the window of the U.K. FAAM aircraft, taken during sampling on 2 April 2012. Photo: Mathias Lanoisellé (RHUL group). (lower right) 30 March onboard CHplot for measurements 5 nautical miles (9.26 km) downwind of the leak; peaks are for measurements at different altitudes—from left 37, 153, and 233 m (see Lee, Mobbs, et al.,). (upper right) Keeling plot of air samples from Elgin methane plume. Inset—enlarged detail. δ‐42.3 ± 0.7 ‰. (2σ error: geometric mean regression). Measurements of CHand δat RHUL (Lee, Mobbs, et al.,).

Fine detail of road circuit around a proposed U.K. shale gas drilling site (Fylde, Lancashire, UK) showing the main sources and their isotopic signatures, and the excess methane averaged across 18 days of surveys and aggregated into 10 m × 10 m grid squares. Farm δ 13 C CH4 signatures are close to −60‰, manure piles −50‰, and gas leaks −40‰. RHUL results.

Case study A: attribution of sources around Fylde, UK . Source attribution by isotopic measurements was used to identify the emissions measured around Fylde, Lancashire, northern England, UK (Figures 7 and 8 ). The measurement, reported here, was carried out as a baseline survey around a proposed site for gas extraction by fracking. Gas sources are identified by their methane/ethane ratios (Figure 9 ) and their different isotopic signatures. Farm signatures have δ 13 C CH4 close to −60‰, manure piles −50‰, and gas leaks −40‰. In a similar study in the Vale of Pickering, also in northern England, gas leaks from the natural gas offtake station had δ 13 C CH4 of −42‰, while farm cow manure piles had δ 13 C CH4 of around −50‰, and methane from the cattle in the barns themselves and their waste had a combined δ 13 C CH4 ‐59‰. However, methane from the local landfill had δ 13 C CH4 of −57‰, closely similar to methane from cattle eating mixed C3/C4 fodder (e.g., grass plus maize). Thus, wind direction is needed to discriminate between landfill emissions and cattle breath.

In these studies, for each emission source type, the isotopic signature and/or ethane:methane ratio must be known for the specific location being investigated. However, both isotopic signatures and ethane:methane ratios vary widely depending on the source, whether gas basin or farm location, and in some cases significant variation in these signatures can occur very locally.

Keeling plots from (a) gas works in Staines (UK), (b) landfill (UK), and (c) three coal mines in the United Kingdom. Errors on theaxis are within 0.05 ‰ and on theaxis within 0.0001 ppmand are not noticeable on the graph. From Zazzeri et al. ().

Zazzeri et al. ( 2015 ) used Keeling plots to characterize sources, using bag samples. Figure 6 a is from a landfill at Mucking in SE England, emitting methane with δ 13 C CH4 ‐56.1 ± 0.5‰ (2 SD) while the Keeling plot in Figure 6 b is from a gas works emitting methane with δ 13 C CH4 ‐36.3 ± 0.3‰. The results show that these two sources are readily distinguishable. In contrast, Figure 6 c shows results from three coal mines. The δ 13 C CH4 intercept from the Thoresby mine in Nottinghamshire, England, was −51.2 ± 0.3‰, markedly different from the methane emitted by the Aberpergwym and Unity mines, which were neighbors 1.2 km apart in the South Wales coalfield. Emissions from the latter two were isotopically similar, with signatures around δ 13 C CH4 ‐32‰ that are just within error and very different from the first mine. Intrinsically methane from these two coalmine emission sources may be separable, with a high enough sample population and also using careful directional measurement by a mobile system. Thus, these results show that even with closely similar sources, such as neighboring coal mines or gaswells, it may be routinely possible to identify emissions by off‐site surface measurement from public access roads. A similar methodology has been used for CO 2 by Domenikos et al. ( 2019 ), while Fries et al. ( 2018 ) used carbon (δ 13 C CH4 ) and hydrogen (δ 2 H CH4 ) stable isotopic determination to distinguish between biogenic methane from sewer gas and thermogenic methane from leaking natural gas pipelines in Cincinnati, Ohio.

This result from Los Angeles was in contrast to the older results of Lowry et al. ( 2001 ), also using stable isotopes, who found a major source of methane emission in the London area was landfills. Lowry et al. ( 2001 ) were reporting on London around the millennium: More recently, much better landfill practice has led to lower U.K. landfill emissions (UK NIR, 2019 ). Currently, Royal Holloway laboratory observations suggest leaking natural gas is the largest contributor to methane emissions around Egham, in western London. This is because all the landfills west of London in the 2001 study have now closed, and gas leaks now predominate.

Townsend‐Small et al. ( 2012 ) used isotopic ratios to interpret inputs to air samples collected across the Los Angeles megacity. They used vehicles to access a variety of sources including oil well fields, oil refineries, power plants, traffic, cattle and cattle feedlots, landfills, and sewage facilities. In addition, they used the geography of the region to collect air from a local mountain (1,707 m above sea level). Methane in the ambient air samples was then analyzed for δ 13 C CH4 and δD CH4 as well as for Δ 14 C CH4 . They found that in Los Angeles the main source of methane emission was leakage of gas from fossil fuel use, including geological sources, natural gas pipelines, oil refining, and power plants. In the Los Angeles study, δD CH4 was the strongest indicator of methane source.

In ambient air relatively remote from sources, off‐line lab‐based isotope ratio mass spectrometry is required to reach the needed analytical precision, by analyzing air samples taken at the study sites. This is because typically it is necessary to determine δ 13 C CH4 to a precision of 0.05‰ or better (Fisher et al., 2006 ) in ambient air masses. Adequate precision is currently achievable by vehicle‐mounted traveling optical analyzers only when very close to source in methane‐enriched air. These instruments typically have isotopic precisions of around 2–3‰, though improved precision can be attained, for example, if the plume can be sampled by AirCore (Karion et al., 2010 ; Rella et al., 2015 ) for isotopic playback. An alternative approach (Lu et al., 2019 ) that shows promise is to use a moving Miller‐Tans analysis (Miller & Tans, 2003 ) of time series data to characterize the population distribution of the isotopic signature for a given source.

Isotopic characterization demands high‐precision measurement of the 12 C: 13 C ratios (expressed as δ 13 C CH4 ) of enhanced and background methane in ambient air masses (Keeling, 1958 ; Pataki et al., 2003 ; Zazzeri et al., 2017 ). In settings where many sources can be geographically closely superposed, isotopic signatures are powerful discriminators between sources. To take an extreme example, an aircraft survey found a methane‐rich plume downwind of the U.K.'s East Anglian offshore gasfields, but isotopic study showed that a significant part of the methane anomaly was not from the gas platforms but had blown there from remote on‐land sources (Cain et al., 2017 ). They found a significant methane anomaly in a plume between 25 and 50 km wide, with elevated methane at 70 to 100 ppb above the background, which was detected during a low altitude (100 masl) overflight over the Leman field's gas platforms in the southern North Sea, off the coast of Norfolk, England. The methane plume was in close proximity above large gas platforms. Air samples were taken and surprisingly showed much of the methane anomaly had an isotopic signature δ 13 C CH4 of −55‰, a characteristic ratio for land sources such as cattle or landfills, and not the −32‰ signature of these gasfields, showing that much of the source was not the gas platforms.

Using an aircraft, Peischl et al. ( 2015 ) measured multiple species in the planetary boundary layer over the major U.S. Haynesville, Fayetteville, and Marcellus gasfields and showed the ratios between methane and other alkanes in the regional atmospheric increment were the same as in local natural gas, thereby proving the increment came from the gasfields. Also, in studies of five U.S. gasfields, Peischl et al. ( 2018 ) used methane/ethane ratios to demonstrate emissions were from the gasfields and not from agricultural sources, given similar atmospheric C 2 H 6 to CH 4 enhancement ratios and the composition of raw natural gas withdrawn from the region. However, Lan et al. ( 2019 ) found an increasing trend in ethane/methane emission ratios in U.S. sources, which may have led to significant overestimation of methane emissions in some studies.

In Los Angeles, Wennberg et al. ( 2012 ) observed the ambient air observations of methane, ethane, and carbon monoxide, in ambient air, and also measured ethane/methane enhancement ratios in the natural gas supply. Assuming nearly all the ethane came from fugitive emissions from the natural gas supply, they were able to attribute most of the excess methane measured in regional air to losses from the gas system, which may have been losing 2.5–6% of its gas to the atmosphere. Similarly, Smith et al. ( 2015 ) used ethane‐to‐methane correlations to quantify the fraction of CH 4 emissions derived from fossil and microbial sources in the Barnett Shale in Texas.

Source identification depends on source flux, temperature, mixing layer height, wind speed, and wind direction. The problem is complex (Lowry et al., 2019 ). Often, major emissions from different source types are closely juxtaposed. Facilities like waste handling and sewage treatment plants tend to be colocated in industrial districts, often with a gas facility, a landfill, and perhaps an incinerator close by, surrounded in many cities by dense suburbs. Sewage treatment plants can be located next to gas distribution installations, or nearby landfills can emit gas close to waste burning facilities. In Australia, the United States, and Canada, gasfields host major cattle feedlots, in some cases sharing water use and capable of emitting more methane than the gas wells. In these settings, source attribution can be challenging. There are so many candidates for methane emissions. Where multiply overlaid sources occur, help can come from of key identifiers such as the presence or absence of associated species such as ethane, or isotopic signatures (Lowry et al., 2019 ). Methane/ethane ratios are powerful discriminants between separate source signals, enabling identification of specific emitters (e.g., Barkley et al., 2019 ; Feinberg et al., 2018 ; Kille et al., 2019 ; Mielke‐Maday et al., 2019 ). Gas leaks contain significant amounts of ethane, while cows do not breathe out much. δ 13 C CH4 studies are powerful tools in pin‐pointing and segregating sources where urban land planning has produced clusters of large but distinct point sources.

A different satellite perspective was taken by Elvidge et al. ( 2009 , 2016 ) who used low‐light imaging data to estimate global natural gas flaring. This major source of both CO 2 and methane emissions remains poorly quantified, especially in regions like Nigeria and Russia. Elvidge et al (2009) showed that in both these regions the flaring efficiency markedly improved over the 2005‐2008 period. More recently, Deetz and Vogel ( 2017 ) used visible infrared imagery to show continued decrease in Nigerian emissions of CO 2 from flaring. If it is assumed that as flaring reduced, methane emissions also declined, this methodology becomes valuable in regional inventory construction.

Finer detail is offered by instruments with pixel resolutions varying from 1–10 km down to as fine as 50 × 50‐m resolution. MethaneSAT, with high spatial resolution and very high precision, is designed to monitor emissions from about 50 oil and gas production regions accounting for about 80% of global production. It will measure methane and carbon dioxide over a ~200km wide swathe (Wofsy & Hamburg, 2019 ).The high‐resolution GHGSat microsatellites designed for greenhouse gas detection, observing methane columns over selected 10 × 10‐km locations, give high effective pixel resolution of 50 × 50 m and 1–5% precision (Jacob et al., 2016 ; Varon et al., 2018 ). Next‐generation versions with excellent spatial resolution and improvement in point source detection better than ~10 kt/yr (0.01 Tg/yr) should be valuable where likely emission loci are already known, for example, in identifying superemitters in gasfields, and in targeting locations of major leaks on gas industry sites.

A powerful example of the usefulness of satellite detection was given by Pandey et al. ( 2019 ) who used TROPOMI observations to discover a very large and hitherto little reported methane emission in the U.S. state of Ohio in February–March 2018. The blowout continued for 20 days, and was observed by TROPOMI on the thirteenth day, when emissions had likely declined from an initial peak rate. With an atmospheric tracer transport simulation, Pandey et al. investigated the methane plume dispersion and used both mass balance and cross‐section flux approaches to quantify the emission rate. They found an emission rate of 120 ± 32 metric tons per hour, which was twice that of the much better reported Aliso Canyon event (Conley et al., 2016 and see below) and that the total methane emission in this little‐discussed Ohio event was perhaps a quarter of Ohio's annual emission. This case demonstrates the power of satellite observation and its usefulness in tracking underreported events.

Recent instruments are considerably advancing the power of satellite observation to help in locating emissions, with better resolution and more frequent observation (Sheng, Jacob, Maasakkers, et al., 2018 ). Turner et al. ( 2018 ) investigated the capabilities of different satellite observing systems. They found the TROPOMI instrument, with 7 × 7‐km pixels, 11‐ppb single‐retrieval column‐weighted precision, and daily frequency (Veefkind et al., 2012 ), should be capable of giving useful information on a spatial resolution of about 30 km. The planned GeoCARB satellite, to be launched in the early 2020s, should give even better resolution (2.7 × 3 km) and precision (4 ppb), making it potentially capable of usefully studying methane emissions from large urban centers such as Shanghai (O'Brien et al., 2016 ), although aerosol pollution will pose a major problem in retrieving accurate methane mixing ratios from regions like China, India, and tropical Africa in biomass burning season. Cusworth et al. ( 2018 ) showed that TROPOMI and GeoCARB will be successful in locating major sources of emissions in spread‐out gasfields but will be less successful where gas installations are densely packed.

However, when used in the context of other inputs and chemical transport modeling, satellite results can be helpful in targeting reduction measures. Specific local enhancements can be detected using GOSAT observations, as shown by Sheng, Jacob, Turner, et al. ( 2018 ), who used satellite results, together with a gridded inventory of emissions (Maasakkers et al., 2016 ), to infer trends in North American emissions from different source types (oil and gas, livestock, and wetlands). In an influential study, Kort et al. ( 2014 ) coupled satellite observation with ground‐based validation to identify the largest satellite‐detected anomaly in the United States within the Four Corners region in the Southwest United States (particularly coalbed extraction), which they showed to have emissions up to 10% of the total U.S. methane emission inventory. In another example, Wecht et al. ( 2014 ) used satellite and aircraft observations, coupled with the local inventory of emissions, to constrain and attribute the geographical distribution of emissions in California. On an even larger scale, Miller et al. ( 2019 ) used observations from the Japanese GOSAT satellite to assess China's methane emissions.

Many earlier studies used the European MIPAS and SCIAMACHY instruments on ENVISAT (Bergamaschi et al., 2013 ; Frankenberg et al., 2006 ; Houweling et al., 2014 ), which also retrieved CO 2 in the 1.65 μm band, and other IR‐active gases, with a coarse pixel resolution. More recently, the Japanese GOSAT (Hu et al., 2018 ; Kuze et al., 2009 ) has provided global mapping of methane, albeit with averaging kernels in the midtroposphere and limited sensitivity to methane in the lower troposphere (e.g., Lange & Landgraf, 2018 ). Although SCIAMACHY (2002–2012) and GOSAT (launched 2009) were capable of identifying regionally important high column measurement over hotspots, the spatial and boundary layer resolution was poor compared to in situ and aircraft mapping, and complicated by biases and poor accuracy of satellite measurement. Moreover, it is not easy to obtain flux quantification from satellite observations. Thus, for the purposes of reducing emissions, satellite observation to date has been of general but not specific value.

By measuring solar backscatter in the shortwave infrared, satellites offer rapid and powerful methods of locating methane emissions (Jacob et al., 2016 ). This is potentially important in the land tropics where in situ observations are typically few and infrequent. Although the utility of satellite observation has hitherto been limited by low measurement frequency, retrieval accuracy, and vertical and pixel spatial resolution, newer satellites, which include commercial satellites that offer pixel‐scene resolutions of the order of a few kilometers and greater sensitivities to the tropospheric column.

Aircraft are useful in measuring agricultural emissions also. Using a light aircraft flying at 50–500 m above ground, Hiller et al. ( 2014 ) tested the Swiss methane emissions inventory over an agricultural valley in rural Switzerland. They measured methane mole fractions during the flights and then used two approaches, eddy flux covariance and boundary layer budgeting, to assess fluxes. Flux estimates were higher than the Swiss national inventory, suggesting Swiss agricultural emissions had been underestimated. Studying a powerful, more focused source, Hacker et al. ( 2016 ) used a low flying aircraft over an Australian feedlot with 17,000 cattle. The aircraft was equipped with quantum cascade laser analyzers, and high resolution wind turbulence measurement. Precision was adequate to apportion emissions qualitatively to individual rows of cattle pens, effluent ponds and manure piles, and elevated methane mixing ratios were detected as far as 25 km from the feedlot.

Karion et al. ( 2013 , 2015 ) and Caulton et al. ( 2014 ) used aircraft‐based measurements to map out emissions from unconventional oil and gas fields in the United States, while Peischl et al. ( 2015 , 2018 ) and Schwietzke et al. ( 2017 ) used aircraft equipped with a high‐precision analyzer to study several major shale gas fields. Similarly, in the Netherlands, Yacovitch et al. ( 2018 ) used a light aircraft equipped with a high‐precision gas analyzer and flask sampling capability to determine emissions from the giant Groningen gasfield and the heavily agricultural surrounding area. They demonstrated that production volume from a specific production site is not a good guide to emissions from that site and that applying inventory emission factors (i.e., production‐based weighting) does a poor job in assessing emissions.

Piloted aircraft capable of carrying high‐precision spectroscopic instruments are still needed for precise measurement of regional fluxes and major emissions sources. To assess emissions from further afield, where mixing with ambient air has taken place and the incremental methane signal is small, for example, some kilometers downstream from a coal mine or large cattle feedlot, or to characterize bulk emissions from a city or a major gasfield, high‐precision optical cavity‐based fast methane analyzers are needed. Currently, instrument and battery packages typically weigh >10 kg for mole fraction precision better than 1 ppb, and the required battery capacity for a useful survey is high (of the order of 100 amp‐hour). These are too heavy to fly on cheap drones and require commercially available UAVs with a takeoff weights typically >20 kg. Moreover, as mentioned above, drones typically have a height limit in most jurisdictions, depending on airspace classification.

In most locations, UAVs are permitted a maximum altitude of 100–200 m, confining them within the lower part of the boundary layer. This is sufficient for proximal surveys on sources, but in most countries higher altitudes are barred to drones and can only be accessed by aircraft. It should be noted however that in many tropical nations UAV restrictions are extremely severe and in some cases wholly barred, particularly near military bases. Moreover, in nations like the United Kingdom legal frameworks for UAV operation are increasingly complex and obtaining permits may become challenging to small academic teams, especially close to sensitive installations such as gas industry facilities, thus may limit the usefulness of UAV techniques for leak studies, and especially the development of novel applications by university research.

On Ascension Island in the equatorial South Atlantic, Brownlow et al. ( 2017 ) and Greatwood et al. ( 2017 ) report bag‐sampling air from as high as 2,700 m in the midtroposphere, well above the height of the local South Atlantic Trade Wind Inversion at about 1,200–1,500 m that defines the top of the marine boundary trade winds. This allowed Brownlow et al. ( 2017 ) to sample air that had traveled from the otherwise difficult‐to‐access Congo basin in equatorial Africa, a potential future tool for validating national and regional emissions inventories.

For isotopic measurements, drones can be used to take grab samples of air in Tedlar or aluminum foil bags (Figure 5 ) to be analyzed in a laboratory. These bags only weigh a few grams and can be carried by small quadcopters. Aircores, which are air samples collected in coiled tubes for later analysis in the laboratory, can also be used to sample air with UAVs (Andersen et al., 2018 ).

UAV surveys are useful in large‐scale screening of gas infrastructure for major point sources (e.g., over a spread‐out gasfield or landfill site) to find and quantify large leaks quickly (Allen et al., 2019 ; Barchyn et al., 2017 ; Yang et al., 2018 ). Many leaks are episodic “one‐offs”—the result of errors such as faulty or forgotten valves, ill‐fitted connections or equipment failure. Sustained drone patrols, carrying low‐cost sensors close to known gas systems, can identify these transient events quickly and cheaply. Such systems have great potential and are likely soon to become a high priority for industry and regulatory bodies (e.g., by the U.K. Environment Agency, see Allen et al., 2018 ). As equipment costs are low, it should be possible and inexpensive to ensure each gas installation has frequent (daily or potentially multiple times a day) routine inspection by automated drones. This may even be profitable, both in terms of less leakage and in safety improvements, averting the risk of major problems and reducing insurance costs.

Recently, more expensive midinfrared lasers have been used, exploiting methane's strong absorption band around 3.3 μm, and have demonstrated performance in the 1‐ to 10‐ppb level (Golston et al., 2017 ; Shah et al., 2019 ) when mounted on drones weighing 5 kg or less. The technology is changing rapidly. Better than 10 ppb (at 1 Hz) precision instrumentation is now available and in use on commercially available UAV platforms (Shah et al., 2019 ), and 1 ppm precision is likely soon. Further marked improvement in the precision, price, and availability of UAV‐mounted sensors should be expected in the near future. Small drone‐carried sensors have much promise for safe close‐up mapping of air with high methane mole fractions around major leaks, for example, in mapping the 3‐D distribution of a plume from a leaking gas installation, and hence quantifying the flux (Emran et al., 2017 ).

Small drones can carry instruments to map methane plumes close to sources and use coincident wind measurements to derive flux using a range of approaches including Gaussian plume modeling and mass balancing. Already it is possible to use these inexpensive light sensors to patrol gas installations. For typical small quadcopter and octocopter drones, the lifting ability is sufficient to carry small sensors able to achieve methane measurement to moderate precision. Most low‐cost low‐weight sensors are capable of measuring methane to 2‐ppm precision (Collier‐Oxandale et al., 2018 ; Shah et al., 2019 ), which should be capable of detecting significant emissions around known candidate sites. However the precision and accuracy of lightweight small sensors is currently poor compared to vehicle‐mounted or static ground‐mounted instruments. Larger UAVs using inexpensive lasers in the near infrared can achieve better precision, to ~100 ppb.

Drones (UAVs) offer powerful opportunities for rapid identification and hence reduction of concentrated methane leaks (Fox et al., 2019 ). Aerial measurement and sampling have the potential to map out emission plumes in three spatial dimensions, as well as time. Moreover, if the drones carry anemometers, eddy flux calculations may be carried out to determine emission fluxes.

Mapping by mobile analyzer is rapid, and field campaign costs can be modest, apart from the capital costs of the instrument and vehicle. In the Royal Holloway laboratory, located SW of the London urban area, Greater London's leaks are being mapped in the duration of a PhD thesis. Hong Kong sources were mapped within a 5‐day working week during a period of steady methane background set by prevailing marine SE Monsoon winds, arriving from the South China Sea. Such campaigns are not in themselves quantitative, nor do they characterize sources as the surveys merely locate major “hot spots”. But in most cases that is enough. Emissions sources can be pinpointed quickly and then remediation or mitigation can be considered. In some cases, halting emissions can be profitable (e.g., from broken pipes or leaking production wells). However, in other cases, where leaks do not pose a significant safety risk, repair may have low priority. In such cases, regulatory intervention may be needed: Thus, there is a strong argument that regulatory authorities should have the capacity for independent discovery of leaks and the power to impose financial penalties.

Helfter et al. ( 2019 ) used measurements from a North Sea ferry to estimate methane emissions from the United Kingdom and Irish republic. They mounted a cavity‐ringdown analyzer on the bow of a commercial freight ferry routinely sailing between Scotland and Belgium and compared the results with measurements made at Mace Head on the west coast of Ireland, using a mass balance approach. They found that estimated annual methane emission from the study region was 2.55 ± 0.48 Tg. This result is comparable with the 2.29 Tg/yr total of estimates submitted to the UNFCCC.

The instruments map out methane mole fraction in the ambient air around the vehicle or above the vehicle's roof. Assuming a steady background air flow, the local increment over background can then be measured while the vehicle moves at speeds that can be up to ~80 km/hr, though typically optimized for plume capture at 20–30 km/hr. Methane peaks identify emission sources: then the sources can be located and mapped carefully using local access roads. Similarly, boats can be used to map emissions from offshore gas platforms (Riddick et al., 2019 ), though here the problem is that the emissions may be from locations such as offshore platform flares that are above the height of the boat's mast intake.

Continuously measuring GPS‐linked mobile systems operate with anemometers to measure wind directions and speeds. As roads typically occur within 100 m to 1 km upwind and downwind of industrial, landfill, and cattle feedlot sources, this allows accurate location of emissions. Thus, in many cases biogas and broad landfill emission peaks can be separated by plume shape and wind direction analysis. In many cases, the likely source can immediately be pinpointed. Figure 4 illustrates a typical vehicle‐mounted system. The costs of vehicle‐mounted campaigns are low. The instrument and associated equipment costs can be below U.S.$100,000. Common SUV vehicles can easily be adapted for the task. They can also deploy air hoses on pop‐up balloons, to take air samples (e.g., for isotopic analysis) (Steiger et al., 2015 ).

It is becoming increasingly inexpensive and rapid to monitor leaks in both production gasfields and urban reticulation systems by routine drive‐by surveys. Rapid precise mapping of methane emissions is now increasingly possible with offsite systems coupled to real‐time global positioning system (GPS) location (Bamberger et al., 2014 ; Phillips et al., 2013 , Rella et al., 2015 , von Fischer et al., 2017 ). Such instruments can be deployed at remote places, and can be mounted on small domestic SUV vehicles (Zavala‐Araiza et al., 2017 ; Zazzeri et al., 2015 ). Vehicle surveys can be supplemented by flying small unmanned aerial vehicles (UAVs) equipped with lightweight sensors (see section 2.2.2 below).

In contrast to camera detection, mobile instruments that map by optical spectroscopy are much more sensitive. They can detect smaller leaks, with good precision and flux quantification (Weller et al., 2018 ) downwind from suspected sources. In the past decade, cavity‐based instruments have radically improved the precision and mobility of in‐the‐field measurement of methane and other associated species (Baer et al., 2012 ; Caulton et al., 2018 ; Crosson, 2008 ). In particular, compared to older gas chromatography, with its low portability and calibration requirements, modern cavity‐enhanced absorption spectrometry instruments have much better precision, robust portability, and more stable calibration in challenging environments (von Fischer et al., 2017 ; Zazzeri et al., 2015 ). Other techniques are also available for methane measurement—for example quantum cascade lasers and open path analyzers (McDermitt et al., 2011 ; Nelson et al., 2004 ). Other species such as ethane can be measured simultaneously, to differentiate between gas supply leaks and sewer emissions. Isotopic analysis can also be used to determine the nature of the source (e.g., gas leaks or sewage pipes) (Fries et al., 2018 ; Lowry et al., 2019 , 2009 ; Phillips et al., 2013 ).

Much progress is being made in leak detection and repair. The U.S. ARPA‐E‐MONITOR program ( https://arpa‐e.energy.gov/?q=program‐projects/MONITOR ) has the ambitious goal to “ detect and measure methane leaks as small as 1 ton per year from a site 10m x 10m in area with a certainty that would allow 90% reduction in methane loss for an annual site cost of $3,000 .” [1 U.S. ton = 907.1 kg]. If this is achieved, it will radically improve our ability to cut emissions not only in the United States but worldwide. Currently, much leak detection in gasfields is by optical gas imaging using passive infrared cameras. For example, over the Bakken Formation in the Williston Basin of North Dakota, USA, Englander et al. ( 2018 ) used helicopter‐borne infrared optical gas imaging to survey about 1,000 well pads, in conjunction with use of a light aircraft to quantify fluxes in a randomly selected sample of 33 plumes. However, the effectiveness of cameras depends on the backdrop, and their usefulness declines with distance, a problem because it limits their usefulness in surprise near‐ but off‐site drive‐by inspections by regulatory bodies. Cameras are effective close to large known sources, such as superemitters, and are excellent for detecting large leaks at gasfields from distances of 10–100 m (Ravikumar et al., 2016 ) and for big emissions over leaking street gas mains, but cameras are less able to find abundant smaller leaks in gas collection, transport and urban distribution networks.

Rapid technical progress is being made, both in pinpointing emissions more accurately and quickly, and in finding smaller point sources. Various vehicle, drone, and aircraft‐mounted technologies can be used. Although most of the simpler current technologies are still only able to provide rudimentary information on flux quantification, they are generally effective in finding leaks and suitable for localization of emission sources (Ravikumar et al., 2019 ).

In this synopsis for Reviews of Geophysics we focus on essentially geophysical methods to reduce emissions. Sustained reduction of the atmospheric methane burden will almost certainly demand much wider actions, both in management of ruminants and also perhaps societal change including demand‐reduction for food derived from ruminants. There is a broad literature on these major topics. The focus here is tactical: what can be done now. The wider strategic issues are briefly discussed later (section 8 ).

Removal of emissions from the air, after they have been emitted, is a major part of the CO 2 carbon capture and storage discussion but has been little discussed for methane. This enquiry is a task that deserves attention and technological development.

Even if major emissions are correctly identified, located, and quantified, then targeting the most cost‐effective ways to reduce emissions is not simple. For example, Kuwait has a major hydrocarbon industry, but field campaigns carried out there by the Royal Holloway (RHUL) group there suggest, counterintuitively, that cutting Kuwait's landfill methane output (Figure 3 ) may be a cost‐effective first‐choice target for cutting overall national emissions (al‐Shalaan, 2019 ), in part perhaps because Kuwait's oil industry already operates a program for emissions reduction.

Accurate quantification of emissions is necessary if the most cost‐effective reduction strategies are to be targeted, but quantification of emission flux in plumes remains an imprecise art. Kang et al. ( 2016 ) showed the disproportionate impact of major gasfield leaks—the so‐called “superemitters.” In the United States Omara et al. ( 2018 ) found the top 5% of sites accounted for 50% of cumulative emissions. Identifying, quantifying, and stopping these superleaks is the first step in emission reduction. More generally, jurisdictions need to quantify, declare, and verify emission declarations at all levels, from local emitter to nation state, but this remains very limited outside Europe (e.g., see Bergamaschi et al., 2018 ) and North America.

Source identification can be a complex task: For example, very different major sources can be closely colocated, such as cattle feedlots within gasfields, or urban landfills near biodigesters, sewage facilities, and gas distribution plants. Industrial area planning and land use often means that such facilities are located near to one another. Urban source types may also be multiple, yet geographically close. Attribution of emissions to specific sources thus can be a challenging task, demanding high‐precision measurement of methane and also winds and boundary layer mixing heights, and often isotopic analysis as well as measurement of other proxy tracer species such as ethane or carbon dioxide (Allen, 2016 ).

Some of these tasks are straightforward but most are not simple. For example, location of emissions, especially major point sources, should intuitively be expected to be an easy task but in practice may be complex as what initially appears to be a point source may actually comprise many small subsources. For example, in a gasfield, leakage from producing wells may be relatively minor, while major emissions may be discovered from unexpected sites such as water processing facilities (e.g., Iverach et al., 2015 ; Kelly et al., 2015 ; Schwietzke, Pétron et al., 2017 ), or forgotten old (legacy) uncapped exploration wells (Day et al., 2015 ).

We need to develop cost‐effective methodologies for cutting emissions and, where they cannot be cut, for removing the emitted methane from air with elevated methane around the sources. Methane reduction is a technical challenge demanding source detection and identification, source‐specific emission flux quantification, and effective reduction methodologies and targeting.

The usefulness of aircraft in major events (see Case Studies A and B above) lends weight to the case for national bodies to maintain aircraft capability as a standing asset for dealing with major air pollution events, also useful in assessing dust clouds from volcanic eruptions, etc.

Schwietzke et al. ( 2018 ) and Englander et al. ( 2018 ) showed that in the U.S. context aerial surveys on light aircraft are an excellent tool, able to locate anomalously highly emitting sources and thus able to target on‐the‐ground vehicle‐based identification and inspection. It is thus becoming feasible to identify leaks where fixing the emission is most cost‐effective (Schwietzke et al., 2018 ). In some cases, it is likely the combination of aerial overpass and ground follow‐up to mitigate emissions can be cost saving and thus profitable to the industry, in other cases, mitigation may have to be driven by regulatory pressure. Aircraft observation of sustained major methane emissions stimulates efforts to reduce unprofitable leakage.

Frequent repeated surveys by light aircraft can track seasonal changes in emission patterns and inform local regulatory bodies of unexpected new inputs or successes in reduction policies. Though derived fluxes can be precise but highly inaccurate, when coupled with technological developments in sampling platforms and new advances in flux inversion modeling such as mass balancing, plume inversion, and gradient flux techniques, calculated fluxes can be used to identify and quantify individual sources and source types (Karion et al., 2019 ).

Aircraft are appropriate in assessing emissions from large conurbations and regions, to verify inventories from wider areas, where there are many sources, such as large gasfields or cities, or wide zones of natural emissions (Miller et al., 2016 ). For example, Cui et al. ( 2019 ) quantified emissions from the Haynesville‐Bossier oil and gasfield by using inverse model calculations driven by aircraft measurements over the gasfield, and showed that emissions were probably larger than estimated in the national inventory. Over the U.S. city of Indianapolis, Cambaliza et al. ( 2015 ) and Heimburger et al. ( 2017 ) used a light aircraft and a mass‐balance approach as well as methane:propane ratios in air samples to quantify urban emissions. O'Shea, Allen, Fleming, et al. ( 2014 ) used upwind and downwind airborne measurements and a mass budget box modeling approach to determine net regional flux of methane for Greater London. In an important study of the Fayetteville shale play, a major U.S. gas producing region, Schwietzke, Pétron et al. ( 2017 ) used aircraft measurement and a mass balance approach, integrating methane mole fraction measurements across plume widths in known wind velocities and direction, to assess methane emissions. The methodology of O'Shea et al. ( 2013 ), O'Shea, Allen, Fleming, et al., 2014 . O'Shea, Allen, Gallagher, et al., 2014 ) and Hartery et al. ( 2018 ), though designed to locate and measure natural emissions, is well suited as a basis for verification of emissions inventories on a regional scale.

As well as cars, bicycles, and child buggies are useful for carrying instruments and batteries, as are backpacks. In a detailed study of emissions around the Munich Oktoberfest in Germany (a major beer festival), Chen et al. ( 2019 ) made in situ measurements by walking and biking around the perimeter of the Oktoberfest. They then applied a Gaussian plume dispersion model to assess emissions.

In practice, each quantification problem is different, and measuring teams have different analytical assets. Ground vehicles are appropriate and widely used for studying locally sourced plumes (e.g., Baillie et al., 2019 ) and can be supplemented with UAV (drone) grab‐bag sampling (Brownlow et al., 2016 ), or by tethered sampling tubes lifted either by UAVs (Allen et al., 2019 ), or pop‐up balloons (Steiger et al., 2015 ). Such approaches can yield 3‐D sampling for local‐scale (e.g., site‐specific) flux retrieval using mass balancing approaches to quantify fluxes accurately. In the cases where a measurement team has a mobile analyzer in a vehicle with a sonic anemometer and GPS location (as in Figure 3 , left), and where low‐level gas plumes can be measured within a few hundred meters of the source, methane emissions can be calculated by driving through the plume. If the source of the emission is known, tracer release can also be an accurate method for assessing methane emissions (Lamb et al., 1995 , 2015 ). Isotopes are particularly helpful in apportioning emissions where various sources are present (Townsend‐Small et al., 2015 ).

While tall tower networks are appropriate as designed for regional budget estimates, their continuous records of methane mole fraction can intrinsically also be used by back‐trajectory analysis to locate major nearby sources, especially if supplemented by air‐sampling for high‐precision isotopic or alkane ratio measurement to differentiate between source types such as landfills or gas leaks. Henne et al. ( 2016 ) validated the Swiss methane inventory using observations from a network of sampling sites in the Swiss plateau and also in the Alps, as the basis of inverse modelling with high spatial resolution <10km over Switzerland. In Australia, Luhur et al. ( 2018 ) used two towers in conjunction with inverse modeling to assess regional methane emissions and quantify fluxes in the Surat Basin from coal seam gas production and processing, coal mining, cattle feedlots, piggeries, wetlands, and other anthropogenic activities. The model domain was 350 × 350 km, which demonstrates the cost‐effectiveness of towers for making a top‐down estimate check of bottom‐up inventories.

In Europe, Bergamaschi et al. ( 2018 ) used data from measurement networks to validate regional emissions, and Wunch et al. ( 2019 ) used long‐term, ground‐based measurements of atmospheric total column methane abundance by remote sensing, to assess methane emissions from the wide northern European region from Poland to France. Their results implied that the European inventories were likely overestimated. The Integrated Carbon Observing System network maintains a long‐term network of stations carrying out high‐precision greenhouse gas measurement ( https://www.icos‐ri.eu/icos‐stations‐network ). For the United Kingdom the GAUGE ( Greenhouse gAs Uk and Global Emissions) Project used a mix of tall tower (Stanley et al., 2018 ; Stavert et al., 2018 ) and other ground‐based measurements, supported by other air‐borne and ship‐borne measurements and satellites with the aim of producing better understanding of U.K. emissions (Palmer et al., 2018 ).

More widely in the United States, Lan et al. ( 2019 ) used long‐term data from both aircraft and tall towers to analyze methane emissions from a variety of sites around the country, including in oil and gas production areas. Encouragingly, this important study found that in the period 2006–2015 the increase in emissions from North American oil and gas industry sources was much smaller than findings in many previous estimates.

In densely populated urban areas, or in major producing regions with dense sources, such as many onshore gasfields, well‐located measurement towers can be used to pinpoint emissions and to assess regional emissions using eddy covariance or optimized Lagrangian‐inversion techniques. This can be applied on any scale, from city to continent or planet. For the US Los Angeles Megacity, Yadav et al. ( 2019 ) took data from a network of 15 in situ sites and used inverse modeling to estimate methane emissions from 3‐km cells, over 4‐day windows, across the megacity area. This method located emissions in enough detail to pinpoint major sources, for example correctly detecting the shutdown of emissions from a major landfill. However, for the very large and highly localized Aliso Canyon leak there were periods when the inversions did not capture fluxes well, especially in unhelpful prevailing wind conditions.

Eddy covariance techniques have long been used to quantify emissions in natural settings (e.g., Detto et al., 2011 ; Rinne et al., 2007 ). In Florence, Italy, Gioli et al. ( 2012 ) sampled from a 3‐m mast that was on a roof 14 m above average roof level and 33 m above ground. Methane fluxes from the footprint area were 189.2 ± 7.0 mg CH 4 ·m −2 ·day −1 and did not show significant temporal variability. Similarly, Helfter et al. ( 2016 ) measured CH 4 , CO 2 , and CO for 3 yr from the BT Tower in central London, 190 m above street level, and from another rooftop site 2 km away, 50 m above street level. Methane emissions in London were similar to Florence, at 197 ± 8 mg CH 4 ·m −2 ·day −1 , but were found to be double inventory assessments and showed moderate seasonality (21% larger in winter). These studies demonstrate the power of the eddy covariance technique in urban settings and its usefulness in testing bottom‐up inventory assessments. In principle, future drones fitted with better sensors and good anemometers will carry out eddy flux calculations directly.

Modeling tools are accessible via U.S. Environmental Protection Agency (EPA) “Other test method” OTM 33a (via https://www3.epa.gov/ttnemc01/prelim/otm33a.pdf ). This is a way of identifying and roughly quantifying emissions from sources located by large spikes in the analyzer record as the vehicle drives past; for example, it is ideal for low‐cost regulatory measurement of emissions from gas wells and pipe leaks.

Caulton et al. ( 2018 ) very successfully used a Gaussian methodology and a hierarchical measurement strategy involving both mobile vehicle‐mounted measurement and small (2–3 m) deployable towers, as well as controlled releases, to quantify emissions from a field of over 1,000 gas wells in Pennsylvania, USA. They clearly demonstrated it is possible reliably to identify and roughly quantify extreme emission sites (i.e., superemitter mitigation targets).

An effective and simple quantification method uses the assumption that the source is a point and that the plume disperses downwind (Brandt et al., 2014 ; Robertson et al., 2017 ; White et al., 1976 ). When methane is emitted from a source, the input is entrained into the moving mass of ambient air. In this moving air mass, the input then disperses both horizontally and vertically, to form a cone of dispersion. If it is valid to assume the gas is well‐mixed in this cone, then the mole fraction of the gas at any nearby point where it can be measured depends on the source strength and height, the wind speed, and the stability of the air mass (Seinfeld & Pandis, 2006 ; Pasquill, 1975 ).

Airborne remote sensing with a visible/infrared imaging spectrometer is extremely powerful providing a regional assessment. In the California Methane Survey, Duren et al. ( 2019 ) measured ground‐reflected solar radiation between 380 and 2,510 nm, from an aircraft at 3‐km altitude, over oil and gas fields, as well as areas with manure and waste management activities. In their campaigns they covered 272,000 facilities, including over 200,000 oil and gas wells. They detected 564 strong point sources emitting methane, especially from landfills. For these distinct sources, measured at 250 facilities, they observed the local methane enhancement and then estimated emission fluxes using wind speed data from weather reanalysis products. They found a population with a heavy tail distribution, with 10% of the point sources responsible for 60% of the emissions. In particular, the largest category of super‐emitters was landfills, especially a subset of superemitters, while dairy farms and oil and gas facilities also contributed.

Pitt et al. ( 2019 ) quantified fluxes using a mass balance approach for the Greater London area using the U.K. FAAM research aircraft ( www.faam.ac.uk ) equipped with a cavity‐enhanced absorption spectrometer. The same aircraft was used by Lee, Mobbs, et al. ( 2018 ), to quantify emission flux from an uncontrolled methane leak from an oil platform in the North Sea. Likewise, O'Shea et al. ( 2013 ) used a cavity‐enhanced absorption spectrometer for airborne measurements over Arctic Scandinavia. With simple box modeling and boundary layer measurement to obtain methane fluxes from wetlands, they found results consistent with parallel in situ chamber‐based measurements on the ground below (O'Shea, Allen, Gallagher et al., 2014 ). With a different approach in flights over England, Pitt et al. ( 2016 ) used an airborne quantum cascade laser absorption spectrometer to measure methane and N 2 O, demonstrating the effectiveness of the method. Moreover, aircraft‐based infrared remote sensing of methane (and other proxy tracers) can offer higher spatial resolution retrieval compared to satellites (e.g., Allen et al., 2014 ).

The aircraft mass balance approach was used by Ren et al. ( 2019 ) to estimate emissions from unconventional gas and oil extraction from the Marcellus Shale in Pennsylvania and Virginia, over a 4,200‐km 2 area. They found emission rates around 0.8% to 1.5% of total gas production, lower than some other studies but consistent with the national inventory assessment. In another study, over major U.S. East Coast cities, Plant et al. ( 2019 ) used aircraft observations of CH 4 /CO 2 ratios and other species to test local emissions inventories. They found that emissions, predominantly attributed to fugitive natural gas, were very substantial and more than twice the recent gridded inventory estimates. Wunch et al. ( 2016 ) used a similar technique with CH 4 /C 2 H 6 ratios to estimate methane emissions in Southern California, another large heavily urbanized urban area. Using historic data, Wunch et al. ( 2016 ) analyzed records of total column abundances for ethane and methane and studied the historic evolution of ethane to methane ratios. They showed that more than half of the excess methane in the Southern California coast basin's air came from natural gas losses.

Proxy tracer (e.g., CO, N 2 O or acetylene, or additional CH 4 ) release methods, using deliberate emission of a tracer gas at a known mass/second rate to validate a dispersion model also offer important techniques to quantify fluxes (Bell et al., 2017 ). The method relies on the assumption that the released tracer gas will behave in the same way as methane emitting from the landfill. The methane emission is then proportional to the ratio of the integrated concentration of the emitted methane to the integrated concentration of the released tracer gas, as measured across the dispersion plume (Scheutz & Kjeldsen, 2019 ). Tracer release can be coupled with spectroscopy to calibrate fluxes from individual sites such as landfill and oil and gas facilities (Ars et al., 2017 ; Foster‐Wittig et al., 2015 ; Rees‐White et al., 2018 ; Roscioli et al., 2015 ; Scheutz et al., 2011 ). Indeed, Scheutz and Kjeldsen ( 2019 ) suggest that the tracer gas dispersion measuring technique should be a core methodology in monitoring to determine emissions.

One of the simplest mass balance quantification methods was used by Lowry et al. ( 2001 ), who studied the London conurbation, which has a detailed CO 2 emissions inventory. They measured the excess CO 2 and CH 4 inputs to urban air, by subtracting the known mixing ratios in contemporary Atlantic background air, and then obtained the methane emissions as a ratio to the inventory CO 2 emissions, assuming the bottom‐up CO 2 inventory to be valid. This allowed them to test the less well constrained London CH 4 emissions inventory for consistency against the better‐constrained CO 2 emissions inventory.

For most sources, fluxes need to be assessed by measuring the air into which the emissions are mixing. In most circumstances it is not possible to enclose sources in chambers, because the sources are unknown, physically too large or disseminated, or otherwise uncapturable. That poses the difficult problem: How does one quantify emissions that are input into moving air masses? The problem is exacerbated, especially in less developed nations, by the lack of access or skills to use sophisticated meteorological products. What may be possible on a flat North American prairie with funding for real‐time local meteorology, good computing skill, and access to sophisticated modeling software may not be feasible for a low‐income community trying to assess fluxes from a landfill located among complex topography under the intertropical convergence zone.

Abandoned wells are relatively straightforward (though not necessarily easy) to locate and once identified may be quickly quantifiable within broad error limits. Although unlined abandoned holes may be difficult to plug, given the cost of both the leak and the explosion danger such super‐emitters pose, they should usually be an obvious target for mitigation. However, many of these historic wells are “orphans” that do not have any permit or current owners, leaving the responsibility for plugging to the state or federal government, which is slow, and mitigation prioritization may not be based on methane emission rate (Townsend‐Small et al., 2016 ).

The simplest way to quantify an emission is to enclose it and measure it directly. Kang et al. ( 2014 ) and Lamb et al. ( 2015 ) placed enclosure chambers around gas leaks and then directly sampled the methane that was captured. Using this methodology, Kang et al. ( 2016 ) demonstrated that 5–8% of annual methane emissions in the U.S. state of Pennsylvania came from abandoned oil and gas wells, many of which were high emitters of methane. Many of the highest emitters are in the Appalachian basin, including Pennsylvania, Ohio, and West Virginia, and they are among the oldest oil and gas wells in the world, some over 100 or 150 yr old (Townsend‐Small et al., 2016 ).

If a big leak occurs, it needs to be fixed. Rapid assessment that a leak is big is essential for safety. Unsafe leaks must be stopped immediately. Thus, in many cases, the only necessary regulatory quantification is the distinction between “large” or “small.” For mitigation however, flux quantification is important. The most obvious challenge is to find those neglected leaks that are large enough to be a significant emission, yet below the “unsafe” or “must‐be‐fixed” threshold. Smaller leaks are of course also important for mitigation, but how should mitigation efforts be targeted? To set priorities for cost‐effective reduction, leak quantification is needed.

4 Practical Emission Reduction and Removal—Tractable emissions

“Tractable” emissions can be defined as those from easily located point sources, for which, once the points are found, mitigation measures can readily be undertaken. In most heavily populated regions the methane source mix is diverse (e.g., Lowry et al., 2001) but widely tractable to mitigation. Such emissions come from many sites—the gas industry, urban wastewater and sewage, landfills, etc. The development of low‐cost methodological advances in leak detection, to identify superemitter leaks, should have a significant impact, especially if better UAV systems become widespread in the near future. It is bad management that leaks gas. It is much in corporate interest to eliminate super‐emitters as they likely cause both loss of profit and also potentially expensive safety risks.

Obvious examples of tractable emissions are known deliberately vented emissions in gas fields and also many emissions in urban distribution systems, both deliberate vents and known leaks that are not large enough to be classed as safety hazards and hence neglected. At the other end of the tractable spectrum are leaks from abandoned installations, and small harder‐to‐find leaks from urban systems. In these cases, deliberate policy is needed to mitigate emissions. However, this is not necessarily very expensive. Kang et al. (2019) found that reducing methane emissions from abandoned oil and gas wells has comparable costs to other greenhouse gas mitigation options.

Typical examples in gasfields (Vaughn et al., 2017; Vaughn et al., 2018) would include methane emitted not only from readily discovered leaks around gaswells but also from deliberate operations such as flaring, venting, and pipe maintenance; unburned gas entrained in compressor exhausts during operation; and also emissions from known processes such as water handling, which can be redesigned or better contained. Episodic venting (Vaughn et al., 2018), which is deliberate, is particularly addressable. Leaks from urban gas networks, landfills, and sewage treatment facility emissions, are widely tractable, as they are known, easily located, and predictable, as are emissions from underground coal mines, where shaft and ventilation outlet locations are well known.

In large gas production and urban distribution systems, discovery and quick approximate quantification is arguably nine tenths of mitigation: Once large sources are known, emissions can generally be ended quickly, often cheaply, even profitably, and with benefit to safety. Both deliberate venting and accidental fugitive releases are readily reducible. Linked to such mitigation measures, strong regulatory efforts can reduce venting and flaring (for U.S. data see https://www.eia.gov/dnav/ng/ng_prod_sum_a_EPG0_VGV_mmcf_a.htm). Flaring and venting gas not only does environmental damage; it also reduces the long‐term energy security of the nation that tolerates the flaring.

4.1 Gasfield Superemitters To cut emissions the most obviously tractable targets are the so‐called “super‐emitters” from the wells, pipelines, pumping stations, and urban distribution networks of natural gas systems. Brandt et al. (2014—see especially their Table S6) showed that in many U.S. gasfields and gas transmission and distribution systems, much of the mass of gas emitted comes from a few major leaks. Similarly, a recent synthesis of methane emissions from the oil and natural gas supply chain indicates that the most emissions come from production wells and that those are dominated by a few major sources (Alvarez et al., 2018). Mineral deposits provide a good analogy. They typically follow Zipf's law of distributions (Guj et al., 2011): in gold mining, there are many small mineral deposits, but few of largest rank, and most of the extractable gold comes from a few very large gold mines. Intuitively, gas leaks would be expected to show similar power law patterns, with a few large leaks and many small, and with a significant proportion of the total amount of leaked gas coming from the few large sources. Detection and monitoring of super‐emitters is rapidly becoming simpler with rapid advances in vehicle‐based sensors (Jackson et al., 2014; Zazzeri et al., 2015), aircraft measurement (Schwietzke, Pétron, et al., 2017) and satellite sensing (Jacob et al., 2016). Satellite observations can now attain high spatial resolution (50 m) at known sites, and are excellent for discovering large emission point sources in remote locations (Varon et al., 2019), but currently remain too sparse to constrain emissions by mapping extensive regions at fine local detail (Turner et al., 2018; Wecht et al., 2014). However, the detection and quantification of a major leakage event in Ohio by the TROPOMI instrument (Pandey et al., 2019; see section 2.2.4) demonstrates that satellite observation is rapidly becoming more able to find and quantify big leaks from superemitters. Measurement by unmanned aerial vehicles (e.g., see the GHGmap project, http://www.geosciencebc.com/projects/2016‐065/), is improving with the development of inexpensive, low‐weight, yet more precise detectors. However, restrictions on UAV flying near “sensitive” facilities like gas installations, may impose significant limits on the future usefulness of UAV‐mounted equipment to monitor facilities, detect leaks and hence mitigate emissions. Aircraft, UAV, and vehicle surveys can now be augmented by direct on‐site continuous monitoring, which is now feasible around production facilities, to watch superemitters and potential super‐emitters across their life cycle, using low cost sensors such as those described by Collier‐Oxandale et al. (2018). Similarly, to achieve higher precision on large offshore gas platforms, automated cavity‐based analyzers can be linked to air inlets distributed around the platform (e.g., an “octopus” design, with air hoses feeding via a central multivalve to the analyzer). In the Barnett Shale gasfield in Texas, superemitting sites accounted for roughly three fourths of total emissions (Zavala‐Araiza, Lyon, Alvarez, Palacios, et al., 2015), which were 90% higher than inventory estimates and constituted 1.5% of natural gas production (Zavala‐Araiza, Lyon, Alvarez, Davis, et al., 2015). Similarly, though with local variation, in four other major U.S. gasfields the bulk of emissions came from 20% of sites (Robertson et al., 2017). Gathering facilities are particularly important sources, especially the leakier installations: in the United States, Mitchell et al. (2015) found 30% of gathering facilities contribute 80% of the total emissions. Abandoned wells also contribute significantly in some regions (Kang et al., 2016; Townsend‐Small et al., 2016). Similar patterns of superemission from abandoned wells and coal‐mine vents were also found in Queensland and New South Wales, Australia (Day et al., 2015; Zazzeri et al., 2016). It will likely need public intervention to locate and cap the multitude of leaking abandoned exploration holes in gas and coal extraction fields. This will at times demand deliberate significant public expenditure for old holes that are “orphans,” with no identifiable or solvent corporate owner. Measurements by Pétron et al. (2012) and Karion et al. (2013) showed very high leakage rates in U.S. shale gasfields: In particular, in the Uintah gasfield in the western USA, Karion et al. (2013) showed emissions were as high as 6.2% to 11.7% of average gas production. These studies have been heavily cited in academic journals and were particularly effective in drawing attention to high loss rates. However, recent gasfield studies elsewhere (e.g., Schwietzke, Pétron, et al., 2017—see above) have found much lower leakage rates, typically well below 2%. Leaks are not the only major cause of gasfield emissions. Schwietzke, Pétron, et al. (2017) and Vaughn et al. (2017) showed that in the Fayetteville Shale, in Arkansas, USA, deliberate manual actions drove major afternoon emission peaks. The western half of the gasfield had much higher emissions (1.8% of production) than the eastern part (0.8%): This was because the water content of the gas was higher in the west, and the higher but episodic emissions could be traced to manual unloading of accumulated water and other liquids. The liquid unloading was by deliberate action for water management. Deliberate venting of unwanted gas is a major source of emission in many oilfields. O'Connell et al. (2019) studied heavy oil wells around Lloydminster, Canada. Over 40% of sampled well pads were emitting detectable methane, and of these, 40% emitted above the venting threshold beyond which mitigation was required by Canadian federal regulations. Similarly, in the Surat Basin, Australia, high‐point vents on the coproduced water distribution lines were found to be important sources of emissions (Day et al., 2015). Major sources of methane emissions from venting and water handling (e.g., Figure 10; Iverach et al., 2015) should be feasible to control, as it should be easy for even ill‐equipped local government authorities to detect and thus regulate them. Good design is generally better than post‐installation mitigation. Thus, when developing new gasfields, consideration should also be given to reducing the number of high‐point vents on the coproduced water distribution pipelines, and where such venting points are required the emissions from these points should be captured. Figure 10 Open in figure viewer PowerPoint (a and b) Methane mole fraction plumes mapped during a mobile survey undertaken by UNSW and RHUL in the Surat Basin in 2014 and presented here. There are 22 similar water holding ponds throughout the Surat Basin. These ponds are required for managing the hundreds of megaliters of daily co‐produced water associated with coal seam gas (CSG) production. Map data: Google DigitalGlobe. Development of effective low‐cost leak detection and monitoring, coupled with targeting superemitters and working with operators to improve operational practices, particularly in water handling, can enable rapid, low‐cost cuts in emissions (Mayfield et al., 2017). Mitchell et al. (2015) pointed to the benefits from simple vigilance: Gas processing plants were staffed and had comparatively low leakage; gathering facilities, typically unstaffed, had larger leaks. In major gasfield regions, even poorly resourced regulatory bodies should be able to afford vehicle‐mounted detection systems of the type shown in Figure 3, which, by imposition of large fines for detected superemitters would likely pay for themselves. Regulation, assisted by tax and penalty nudges incentives, is an obvious first step to reduce leaks and flaring from gasfield sources. In considering mitigation and safety (and perhaps profit), self‐interest can be made to converge with greenhouse good. In some superemitter cases the economics of mitigating very large emissions are so attractive that high leakage simply denotes bad cost control and incompetent corporate leadership. If overall profit margins are, for example, 10%, a small investment to recapture a 10% leak will double profits, and enhance safety. Canada has introduced some controls, intending to cut methane emissions by 40–45% by 2025, but clearly ineffective as yet in the case of the Lloydminster emissions reported by O'Connell et al. (2019). Tyner and Johnson (2018), studying oil production sites in Alberta, Canada, used Monte Carlo simulations to show that a 45% reduction in methane emissions from flaring and venting would, per ton of CO 2 equivalent greenhouse warming, cost between about CAN $2.5 to CAN $−3 (i.e., a profit). Relatively small government tax incentives for mitigation and relatively large penalty disincentives imposed on leakage would likely serve to nudge the mitigation of emissions well beyond a 45% cut. In any nation, even such relatively lax requirements do help cut emissions, with the side impact of slightly enhancing national energy longevity and security of supply. In 2016, as part of the United States' plan to meet its Paris Climate goals, the U.S. EPA introduced the Oil and Natural Gas Sector New Source Performance Standards, with a variety of emissions reductions requirements wells and compressors constructed after September 2015, including regular leak detection and repair for wells and compressor stations. However, EPA has challenged this rule in court and has reduced the amount of monitoring required since the regulation was implemented. The U.S. Bureau of Land Management also implemented a Waste Prevention Rule in 2016 to reduce methane losses from oil and gas wells on federal land, but many aspects of that rule have also been repealed recently. The intention of the 2016 EPA action was to decrease methane emissions by 300,000 short tons (unit from original citation − amount = 0.27 Tg) in 2020, and by 510,000 short tons in 2025 (0.46 Tg), compared to a baseline estimate. In contrast, under the revised 2018 EPA proposal, U.S. emissions were perversely expected to increase by 380,000 tons (0.34Tg) between 2019 and 2025 (Reuters, 2018a): In effect this emits future U.S. energy security to air. Interestingly, this 2018 EPA proposal was opposed by the industry, which is well aware of costs: An Exxon Vice‐President stated “We support maintaining the key elements of the underlying regulation, such as leak detection and repair programs” (Reuters, 2018b).

4.2 Urban Gas Leaks Urban gas leaks are major sources of methane emissions in most long‐urbanized societies (e.g., Lowry et al., 2001; McKain et al., 2015; Zazzeri et al., 2017). Many cities in developed nations have built inventories of estimated methane emissions. However, using aircraft‐mounted observations downwind of major U.S. cities, Plant et al. (2019) found that in current inventories, natural gas emissions were significantly underestimated. Moreover, in many places leak detection is still primarily by human smell. Although methane is odorless, this is effective as a safety measure because urban gas typically has the organosulfur compound mercaptan (CH 4 S) added, which can be detected at a 0.14‐ppb threshold by the human nose (Committee on Acute Exposure Guideline Levels, 2013). In the Boston, USA, urban region, McKain et al. (2015) reported loss rates to the atmosphere of around 2.7%, as a proportion of the gas used. In the U.S. Los Angeles basin, He et al. (2019) found that about 1.4% of commercial and residential gas consumption is released into the air. Such loss rates are expensive, locally potentially dangerous, and an unnecessary greenhouse emission. Yet currently many jurisdictions pay little attention to mitigating urban gas leaks: for example, in the United States, Hopkins et al. (2016) concluded that “current mitigation approaches are absent or ineffective.” Even safety can be neglected. A recent example is the series of explosions and fires in the Merrimack valley, Massachusetts, USA, in September 2018 that killed one person and damaged over a hundred buildings (NTSB, 2018). Surveying the urban distribution network in Boston, USA, Hendrick et al. (2016) found that 15% of leaks surveyed were potentially explosive, and even small leaks could not be disregarded as “safely leaking.” As with gasfields, there is strong evidence that a few major leaks constitute the bulk of emissions from urban distribution networks and that there is a “fat‐tail” distribution of leaks. In Boston, 7% of leak sources contributed 50% of emissions (Hendrick et al., 2016) and in U.S. cities, it is estimated (von Fischer et al., 2017) that repairs to the largest 20% of pipeline leaks would cut losses by half. Leak prioritization by mobile measurement is now rapid, inexpensive, and effective (Weller et al., 2018). Major leaks stand out strongly in vehicle surveys: Their identification does not need sophisticated quantification. Around the Royal Holloway location in outer London, gas governor stations, where pressures are controlled for domestic distribution, are particular point sources of emissions. It is likely that much could be done to cut emissions by identifying such specific “likely to leak” parts of the distribution network as mitigation targets. There are strong grounds (Hausman & Muehlenbachs, 2016) for emphasizing leak reduction over pipeline replacement as a first option in mitigating emissions from a gas distribution system, especially with the marked advances in leak detection. Leak quantification will help optimize mitigation: Identifying and repairing the worst emissions is an attractive and potentially profitable first step. There is indeed evidence that in the U.S. distribution pipeline leaks are being reduced (Lamb et al., 2015). However, it is difficult for gas distribution companies to make the jump between leak detection and leak quantification, and their criteria for repair or replacement may have different priorities (i.e., safety and proximity to structures, or cost, rather than greenhouse gas emissions). Currently, it is possible that widespread practice may be to mitigate methane leaks for safety reasons, but not to be overly concerned about greenhouse implications. In addition to leak reduction, the replacement of century‐old iron pipes is urgently and widely needed in Europe and the United States. This is necessary to cut greenhouse emissions and also to improve safety, by reducing unexpected catastrophic leaks, although the replacement process itself can be hazardous (NTSB, 2018). Robotic autonomous systems offer much (Robotics, 2017). The use of robotic systems removes the need to dig up the pipe to seal it. These systems can travel along pipes, detecting leaks and in principle also sealing many from the inside, including those in “live pipes filled with gas, and potentially giving longer lifetimes to olds iron pipes.” In the United Kingdom, replacement is typically achieved without major excavation by inserting new narrower‐diameter higher‐pressure piping within the older wider‐diameter iron pipes. However, in countries with regulated local gas monopolies any investment in leak reduction increases the capital base of the gas industries and thus increases allowable gas prices. Hence proposals from industry to invest in new pipes may be blocked by regulators. For example, in the past the U.K.'s Office of Gas Supply (Ofgas) resisted environmental objectives (Danby, 1998). To keep prices down, U.K. regulators restricted corporate proposals for investment in leak reduction by pipeline renewal. While this may have been politically advantageous in the short term, this extra leakage would have proportionately slightly shortened the useful life of the UK's domestic North Sea gas supplies, quite apart from the environmental damage.

4.3 “Habitual” Emissions From Domestic and Industrial Facilities “Habitually” high methane air commonly exists in many places, ranging from near vents of domestic boilers and water heaters to the surrounds of industrial gas handling facilities such as compressor stations and gas distribution centers (Figure 11) (Schwietzke, Pétron, et al., 2017; Subramanian et al., 2015; Zavala‐Araiza et al., 2017). In dealing with the high numbers of small leaks and emissions, low‐cost sensors are becoming capable of detecting methane anomalies at the sub‐ppm level (e.g., Collier‐Oxandale et al., 2018). Just as smoke detectors are now ubiquitous, so methane sensors could be widely deployed to identify leaks rapidly. Figure 11 Open in figure viewer PowerPoint Typical source facility: gas processing plant, Hong Kong. Photo: E.G. Nisbet. Domestic heating systems emit methane on ignition because the first gas is only partly combusted. In the U.K. domestic boilers emit when they ignite for house heating and hot water. In some nations (e.g., Australia) “instant” hot water heaters that rely on burning gas are common: they emit methane each time the hot water tap is turned on. When a gas‐burning system such as a domestic boiler or industrial compressor switches on or off, the combustion chamber is vented to prevent explosive build‐up of unburnt gas. In addition, steady‐state burning itself is inherently incomplete. In order to ensure all gas is oxidized and no stray gas leaks out, down to the ambient background level, the amount of oxygen in the system must be increased well above the stoichiometric level. The reduction in overall combustion chamber temperature caused by this excess airflow reduces the boiler efficiency by more than the energy cost of venting unburnt gas. Because unburned gas passing through can be significant when burners are first turned on, an alternative strategy to cut emissions may be to pass initial exhaust gas back into the burning chamber for a second combustion. To mitigate emissions, not all the methane needs to be removed: the challenge is to devise an inexpensive, robust system to reduce at least some of the excess methane in the exhaust. Catalytic removal of a significant fraction of excess methane is feasible. Removing harmful trace gases is compulsory for vehicles: In a comparable way, in future methane removal from exhaust gases may become an attractive and feasible option for stationary gas‐using installations. Catalytic reactors can be autothermal at methane mole fraction as low as 0.06% but for greenhouse mitigation the catalyst would need heating (Su & Agnew, 2006). Noble metal catalysts using noble metals such as Pt, Pd, or Rh are effective (Jiang et al., 2018) and in principle it should be possible to take advantages of economies of scale by using modified vehicle catalytic converters. But costs may be prohibitive. Gosiewski and Pawlaczyk (2014) estimate that 0.5 t of Pd would be needed to mitigate emissions from a single coal mine ventilation shaft. Unfortunately, currently, there has been little research on ultralean methane combustion using these catalysts (Jiang et al., 2018), though palladium‐based zeolite catalysis is possible (Petrov et al., 2018). More generally, the use of zeolite‐metal technologies to remove methane from air is promising (Jackson et al., 2019). Inexpensive metal‐oxide regeneratable catalysts (Buciuman et al., 1999; Döbber et al., 2004) have long been used in laboratory zero‐air generators in reversible flow reactors. Typically, these reactors use catalysts such as hopcalite (a mixture of copper and manganese oxides, sometimes with cobalt or other metals), which work well at moderate temperatures (below 600 °C). These catalysts are typically rapidly made ineffective by water but can be easily regenerated by heating. Luo et al. (2012) showed full combustion of methane at 575 °C using biomorphic CuO‐ZrO 2 catalysts synthesized on cotton templates. There is also great potential for the direct catalytic conversion of methane to methanol at low temperature (~220 °C or less) over zeolites that have been copper‐exchanged (Narsimhan et al., 2016). UV‐photocatalytic removal of methane may also be feasible (de Richter et al., 2017; Graetzel et al., 1989). As methane catalysis only occurs at temperatures above several hundred degrees, a vital consideration in catalytic methane reduction is ensuring that the global warming caused by the extra energy demand needed to operate the catalyst is smaller than that of the mitigated gas, that is, that there is net benefit. For simple estimation the heat capacity of 1 m3 of air is about 1.2 kJ. If a catalytic converter raises the exhaust temperature by 250 °C before venting then the converter must reduce the exhaust methane concentration by well over 10,000 ppm (assuming heating at 0.1 g CO 2 eq per kJ). This should be compared to exhaust concentrations from domestic solid‐fuel boilers of 10–100 ppm (EU Ecodesign Directive 2009/125/EC). One solution to this is to place the catalyst inside the combustion chamber where temperatures are already sufficient for catalysis and exhaust gas is still hot as it passes over. Vaillant and Gastec (1999) have demonstrated such a system (Figure 12) and achieved 0‐ppm exhaust emissions, an improvement even over ambient mole fractions. An alternative catalytic approach is to replicate the natural oxidation of atmospheric CH 4 by OH radicals—this is discussed further in section 6. Finally, it is worth remarking that a more elegant solution to the problem of gas emissions from domestic boilers would be to eliminate the domestic gas boiler and its pipeline supply system entirely. Renewable energy sources such as wind and sunlight are sporadic. Inevitably, in power systems with high proportions of renewable generation, there are episodes of superabundant electricity, for example, at 3 am on a windy night, or on a quiet very sunny afternoon. As decarbonization proceeds, there will be many times of surplus supply at periods of low demand. Moreover, the costs of nuclear power are typically dominated by capital cost, and running costs are comparatively small. Thus, nuclear electricity can also be very inexpensive at 3 a.m. Gas is primarily used either for domestic heating or for electricity generation. In European countries domestic central heating based on circulating hot water is common and wind, and solar power are widely and increasingly used, coupled with a baseload of nuclear power. Here, rather than using on‐demand gas boilers, it is likely to become preferable to use surplus‐period (e.g., postmidnight) electricity to heat water, and then release the heat into the home over the next day and evening. Hot water can easily be stored in an insulated tank for several days without losing much heat. Cold can obviously be stored also: In hot places it may become preferable to operate air conditioning by blowing air over a thermal mass such as ice chilled at a time when electricity was abundant. The intrinsically leaky urban gas pipeline network could be repurposed as cabling conduits both for telecommunication and, as vehicles turn electric, for electric power supply. Closing the domestic gas network and switching to renewable‐generated electric heating may have considerable synergy with the move to battery‐powered electric cars. In the United Kingdom it would benefit long‐term energy security if gas imports were halted and the declining local North Sea gas reserves were retained solely for use in peak‐demand electricity generation, rather than used up for domestic heating.

4.4 Urban Wastewater and Sewage Sewage systems host anaerobic methanogens and thus emit methane at all stages, from house drains to wastewater treatment plants (Liu, Ni, et al., 2015; Guisasola et al., 2008, 2009; Fries et al., 2018). Where biogas is extracted, it may be incompletely combusted. Drains are prone to methane explosions, while methane production in the sewage reduces the readily biodegradable chemical oxygen demand (Guisasola et al., 2008), detrimental to processes in the wastewater treatment plants. Thus, methane emissions from sewage constitute a safety and processing problem, as well as being a greenhouse concern. The wider problem of methane removal from sewers in the long gravity runs from buildings to wastewater treatment plants is potentially addressable by installing small bioreactors at intervals along the length of the sewers, and by extracting methane‐rich air above the sewage for catalytic or soil methane oxidation. Figure 12 Open in figure viewer PowerPoint 1999 Concept for completely catalytic boiler system. The premixed gas air mixture is distributed over a heat exchanger. Modified on a concept from Vaillant and Gastec (). Various techniques can be used to reduce emissions, but at present efforts have focused on the wastewater treatment plants. Here, methane recovery efficiencies over 50% can be achieved, for example, via a submerged underwater bioreactor (Giménez et al., 2012). Floating bioreactors on wastewater settling pools have achieved 67% methane oxidation in agricultural settings (Syed et al. (2017), and the technology may be transferrable to urban wastewater. Matsura et al. (2015) achieved over 99% removal of dissolved methane by a two‐stage system using a downflow hanging sponge reactor to remove methane from an upflow of anaerobic sludge effluents. Similarly, using an upflow anaerobic sludge blanket reactor at ambient temperatures, Bandara et al. (2012) achieved good removal efficiencies of more than 50% in summer, as measured by chemical oxygen demand, though less than 40% in winter. No biogas was evolved. More generally, methane removal in wastewater treatment plants appears to be both feasible and inexpensive, partly self‐financing in both reduction of safety risk and perhaps in providing methane for ancillary power.

4.5 Methane Emissions From Biodigesters and Biogas Generators Biogas and oils are increasingly being produced by anaerobic digestion of various substrates such as domestic and industrial food waste, manure, or sewage sludge. There are significant benefits from this activity, including energy generation that replaces fossil fuel combustion, and mitigation of emission of methane from manure. However, as biogas is typically 50–70% CH 4 and 30–50% CO 2 , if part of the biogas is lost to the air, then there is a risk of cutting CO 2 emissions but increasing net greenhouse warming impact by the emission of methane. The production of renewable energy from biogas plants is growing rapidly worldwide. In the United Kingdom, biogas plant numbers have increased almost 500% in the past 5 yr, with 607 operational biogas plants in 2018, and around 500 new plants being considered for construction (ADBA, 2018). The European Biogas Association anticipates that overall biogas production will be at least 50 × 109 m3 by 2030, capable of providing 2–4% of the EU's electricity needs by displacing methane from fossil fuels (EBA, 2019). Unfortunately, biogas plants can have very high fugitive methane emissions. The emission rate depends on engine construction, plant design, and operation. Emissions can come from water handling facilities, gas engine exhausts, gas flaring, leaks from pipes, biogas upgrading units, tanks, etc., or from deliberate venting (Angelidaki et al., 2018; Duren et al., 2019; Fredenslund et al., 2018; Liebetrau et al., 2018; Samuelsson et al., 2018; Scheutz & Fredenslund, 2019). Some emissions can be single large leaks or long‐lasting bursts from pressure relief valves (Reinelt & Liebetrau, 2019). In particular, Kvist and Aryal (2019) found that water scrubbers and pressure relief valves were especially significant sources of emissions, a finding similar to findings around many unconventional natural gaswells. To 