Emissions of air-borne mercury from five municipal solid waste landfills in Guiyang and Wuhan , China

China disposes of bulk Municipal Solid Waste (MSW) by landfilling, resulting in a large quantity of mercury that enters landfills through waste. A detailed study on atmospheric mercury emissions from MSW landfills in China is necessary to understand mercury behavior from this source. Between 2003 and 2006, mercury airborne emissions through different pathways, as well as mercury speciation in Landfill Gas (LFG) were measured at 5 MSW landfills in Guiyang and Wuhan, China. The results showed that mercury content in the substrate increased the magnitude of mercury emissions, with the highest emission rate measured at the working face and in uncovered waste areas, and the lowest measured near soil covers and vegetated areas. Meteorological parameters, especially solar radiation, influenced the diurnal pattern of mercury surface-air emissions. Total Gaseous Mercury (TGM) in LFG varied from 2.0 to 1406.0 ng m−3, Monomethyl Mercury (MMHg) and Dimethyl Mercury (DMHg) in LFG averaged at 1.93 and 9.21 ng m−3, and accounted for 0.51% and 1.79% of the TGM in the LFG, respectively. Total mercury emitted from the five landfills ranged from 17 to 3300 g yr −1, with the highest from the working face, then soil covering, and finally the vent pipes. Correspondence to: X. Feng (fengxinbin@vip.skleg.cn)


Introduction
Mercury (Hg) emissions from Municipal Solid Waste (MSW) incineration are regarded as one of the most important anthropogenic mercury sources to the atmosphere (US EPA, 1997;van Velzen et al., 2002).However, about 70% of MSW globally are deposited in landfills due to their low cost and low maintenance (OECD, 2001).Other researches showed that landfill also served as a potential atmospheric mercury source (e.g.Lindberg and Price, 1999;Kim et al., 2001), and more importantly, landfill sites were effective bioreactors that converted inorganic mercury into methylmercury, the most toxic form of mercury (Lindberg et al., 2001(Lindberg et al., , 2005a;;Hawkins and Prestbo, 2004;Prestbo et al., 2003).
China is one of the largest MSW producers in the world, generating over 150 million tonnes MSW per year, which accounts for 29% of the global total.Landfill is the main disposal method for MSW in China, treating 68.9 million tonnes at 444 landfills in 2004, accounting for 85.4% of the total amount of treated waste (including landfill, incineration and compost, the rest was not treated properly, Huang et al., 2006;Liu et al., 2007).The area occupied by MSW landfills in China is about 500 km 2 , and the volume of buried MSW has reached up to 664 million m 3 in 2008.Mercury enters the landfill mainly through mercury-containing waste, such as batteries, fluorescent lamps and thermometers (US EPA, 1992).Between 1992 and 1999 in China, 185-802 tonnes of mercury was released in the environment from discarded batteries (Yang et al., 2003).Although Hg content in batteries was lowered since 2001, 153 tonnes of Hg were still used Published by Copernicus Publications on behalf of the European Geosciences Union.  in batteries in 2004 (Jian et al., 2008).In fluorescent lamps and thermometers, about 200 tonnes of Hg are used each year (Hao and Shen, 2006;Shen and Jian, 2004).Because most (over 90%) of the Hg-containing products were not recycled in China (Yu and Li, 2004), they end up in landfills.Hg emissions from other anthropogenic sources in China, such as coal combustion (Wang et al., 2000(Wang et al., , 2010;;Tang et al., 2007) and nonferrous metal smelting (Feng et al., 2004a;Li et al., 2009) have been extensively studied, but little information is known about mercury emissions from landfill sites.Unlike other western countries, there are very few facilities that utilize Landfill Gas (LFG) in China (Huang and He, 2008).As a result, almost all the LFG is emitted into the atmosphere directly, which poses a severe ecological risk.Thus, mercury emissions from landfill sites in China deserve investigation.
In this paper we report the results of mercury emissions from 5 MSW landfills in Guiyang and Wuhan city, China, sampled between 2003 and 2006.Hg emissions were measured from two pathways, i.e., landfill surface and landfill vent pipes, and Hg species in LFG, including total gaseous Hg (TGM), monomethyl Hg (MMHg) and dimethyl Hg (DMHg), were also measured.

Landfills studied
Locations of five investigated landfills are given in Fig. 1, basic information is listed in Table 1, and photos showing the sampling sites are shown in Fig. 2. Three landfills are located in Guiyang (capital of Guizhou province), namely Gao-Yan (G-Y), Da-Zhuan-Wan (D-Z-W) and Xian-Ren-Jiao (X-R-J) landfill, and two in Wuhan (capital of Hubei province), namely Jin-Kou (J-K) and Dai-Shan (D-S) landfill.Due to its karstic landscape, landfills in Guiyang are located in valleys, while landfills in Wuhan are located in flat areas.Guiyang has a population of 3.3 million and produces 2100 tonnes of MSW per day, while Wuhan has a population of 7.8 million and produces 6065 tonnes of MSW per day.The climate of both cities is a typical northern subtropical monsoon.
Of the five landfills, two were sanitary landfills (G-Y and J-K), and three (D-Z-W, X-R-J and D-S) were simple landfills.During the study period, three landfills (G-Y, J-K and D-S) were in active operation, while the others were closed.A passive LFG vent system was installed at G-Y and J-K landfills, discharging LFG directly into the ambient air (see Fig. 2d), while the other three landfills did not employ the vent pipe system.The landfills studied are representative of Chinese landfills, since they include different types (sanitary or simple), different stages (operational or closed), and different surface types.

Sampling and analysis methods
The field sampling campaigns were carried out between 2003 and 2006 at G-Y landfill, while the other 4 landfills were measured in 2004.Monitoring focused on two aspects, namely Hg surface-air flux and Hg speciation.

Hg surface-air flux
Two methods were applied to measure Hg surface-air flux, i.e., a Dynamic Flux Chamber (DFC) method for the nonworking face area (Fig. 2c), and a Gaussian plume model for the working face area, where the MSW was dumped, spread, crushed and covered with soil.The principal of DFC measured in conjunction with an automated mercury vapor analyzer (Tekran 2537A) was described elsewhere (Feng et al., 2005;Wang et al., 2005).Briefly, the DFC was a bottom opened, semi-cylinder shaped quartz glass chamber with a rectangular (20×30 cm) area tight towards the substrate.A turnover time of the air enclosed of was about 10 s.Mercury surface-air flux F Hg was calculated according to Eq. ( 1): with a time resolution of 20 min, where Q is the air flow rate through the chamber, A is the area exposed to the surface as well as c out and c in are the mean Total Gaseous Mercury (TGM) concentration measured in the outlet and inlet air, respectively.The Tekran 2537A is a sensitive analyzer for TGM with a low detection limit (about 0.1 ng m −3 , e.g., Tekran, 1998), and calibrated by its interior mercury vapour source.The chamber blank was 0.5±1.8ng m −2 h −1 (n=77).
The Gaussian plume model (US EPA, 1995) used here is the ISCST3 model (Industrial Source Complex, Short-Term, Version 3), and is expressed as follows, where C(x,y) is the atmospheric pollutant concentration at the horizontal coordinates x, y; Q is the pollutant emission rate from the source; U is the horizontal wind speed; h is the effective height of pollution source; σ y and σ z are horizontal and vertical dispersion coefficient respectively; x, y, z are Cartesian coordinates.
The working face was treated as a ground surface source, thus the Hg emission rate can be deduced from the meteorological parameters, and atmospheric Hg concentration at downwind and upwind working face sites.The dispersion coefficient was calculated through onsite meteorological observations.

Hg speciation in the LFG
Hg speciation in the LFG, including TGM, MMHg and DMHg, was determined at G-Y landfill, while only TGM  was measured at J-K landfill.The sampling and analysis methods for mercury speciation were described by Lindberg et al. (2001) and Bloom et al. (2005).TGM was measured by the Tekran 2537A onsite at 5 min intervals, and MMHg and DMHg were trapped by diluted HCl (0.5% v/v in double deionized water) and Carbotrap TM adsorbent (20/40 mesh, Supelco Inc., Bellefonte, PA), respectively, with sampling time more than 2 hours.Subsequently, sorbents and aliquots employed to trap organic mercury forms were transported to the laboratory, where DMHg and MMHg (the latter after aqueous ethylation) were determined by gas-chromatography separation, thermal reduction to Hg 0 and CVAFS detection (Tekran Model 2500, Canada).Excessive moisture in LFG was removed by passing the air stream through a condenser (an impinger in ice bath equipped with a soda lime desiccant trap).The Carbotrap TM was wrapped with aluminum foil during the sampling and storage period to avoid decomposition of the analyte.MMHg and DMHg in the LFG were calibrated by purging MMHg standard solution (supplied by Cebam Analytical Inc., WA, USA) subjected to ethylation and aliquots of DMHg standard solution (supplied by Jožef Stefan Institute, Ljubljana, Slovenia) respectively onto Carbotrap TM adsorbents, which being analyzed as described above.During each campaign, the field blank samples for MMHg and DMHg were also collected.The method detection limits for MMHg and DMHg were found to be 0.5-0.6 pg (as mercury) absolute or 1.4-1.7 pg m −3 (for 2 h sampling) based on 3 times deviation of the blank samples.

Other parameters
In addition, Hg content, pH and Organic Matter (OM) in the MSW and cover soil of the 5 landfills, TGM in the ambient air above the landfill surface (0.1-2.0 m high), and meteorological parameters, including air/soil temperature, relative humidity, solar radiation, wind speed and wind direction were also monitored to characterize the behavior of Hg emissions from landfill sites.The meteorological parameters were monitored by using a portable weather station (Global Water IIIB, USA).TGM in the ambient air was determined onsite by Tekran 2537A.Hg content in the MSW and cover soil was analyzed by CVAFS detection (Tekran Model 2500, Canada) after aqua regia (HCl+HNO 3 3:1 v/v) digestion.While, pH and organic matter in the MSW and cover soil were measured by a pH meter in a 2.5:1 (v/m) water/solid suspension and potassium dichromate method, respectively.

Hg, pH and OM in the MSW and cover soil
Table 2 includes Hg, pH and OM content in the MSWs and cover soils.Hg in MSW was characterized by significant variability (Guiyang, range: 0.170-46.222mg kg −1 , average: 1.796±7.072mg kg −1 , N =42; Wuhan, range: 0.240-1.271mg kg −1 , average: 0.606±0.349mg kg −1 , N=8).The result for Wuhan was very similar to that reported by Fang and Hong (1988), where Hg concentrations in MSW of Wuhan ranged from 0.13 to 1.53 mg kg −1 and averaged at 0.52 mg kg −1 (N=20).Hg concentrations in cover soils, in general, were more convergent, with averages of 0.310 and 0.058 mg kg −1 for Guiyang and Wuhan, respectively.These values are similar to typical background levels, including 0.222 mg kg −1 for Guiyang (Wang, 2004) and 0.056 mg kg −1 for Wuhan (Fang and Hong, 1988).A few cover soil samples in D-Z-W landfill contained high Hg (3.124-6.527mg kg −1 ), which may be due to unauthorized dumping of MSW that occurred at this landfill after its closure.Hg concentration in MSW was higher compared to cover soils, which possibly reflected higher mercurycontained substances in the MSW.Organic matter content and pH in MSW were also elevated compared to cover soils (Table 2), due to co-mingling of kitchen waste (such as food remnants and leaves, etc.) and coal ash (pH 7.5-12.1,from the domestic cooking and heating), respectively.

TGM in the atmosphere above the landfill
Figure 3a shows the ranges of TGM concentration observed in the ambient air over the landfills studied.The range of TGM was from 1.6 to 473.7 ng m −3 , with averages (Fig. 3b) at different sites ranging from 8.5 to 155.7 ng m −3 .The highest TGM concentrations occurred at the working face and the downwind area for all three landfills (G-Y, J-K and D-S), where TGM was sensitive and proportional to the activities of MSW treating at the working face, as observed at operational landfills in Florida, USA (Lindberg et al., 2005b).Lowest TGM was measured at the closed landfill of X-R-J, where the whole landfill was planted with grass and trees, and this value was close to average TGM concentrations in ambient air in Guiyang (8.4 ng m −3 , from Feng et al., 2004b).Since the landfills were located far from other urban Hg emission sources, the elevation of TGM concentrations in the ambient air was predominantly due to the landfill emissions.

Non-working face areas
Hg surface-air fluxes at the non-working face areas, as determined by the DFC method, are listed in Table 3.The flux indicated large variability from site to site, ranging from −286.2 to 5609.6 ng m −2 h −1 , with highest averages of about 500-600 ng m −2 h −1 for the sporadically uncovered MSW sites (site F6, F10 in Table 3) and the contaminated soil cover area (site F1), while the lowest was observed at soil covers (site F13, F17) and the grass planted area (F18) with average rates about −1 to 20 ng m −2 h −1 .Hg flux was clearly higher during the warm season for the same surface type, such as "uncovered MSW" sites at warm season (F6 and F10) versus cold season (F14 and F15), and "temporary soil cover" sites at warm season (F8 and F9) versus cold season (F12 and F13).The detailed processes of Hg surface-air flux, as well as the concurrent meteorological parameters at each type of surface are illustrated in Figs.4-6.Among all parameters, a strong diel cycle with a daytime maximum was observed.Figure 4 shows the contributions of high Hg content (2.313 mg kg −1 ) in the MSW to the emission rate.Figure 5 shows how vegetation reduced the mercury emission rate, since the mercury surface-air flux at grass planted area of X-R-J was obviously lower than that of two soil covers (with no plantation) at D-Z-W, although under similar weather conditions and with similar mercury contents in the substrate (0.477-0.575 mg kg −1 ) among three sites.At two similar soil cover sites of J-K landfill, Hg flux was several folds higher under sunny conditions compared to cloudy and rainy conditions (Fig. 6).
Compared with other studies, Hg emissions from soil covers were higher than those from several landfills in Florida, USA, which ranged from 1-20 ng m −2 h −1 (Lindberg and Price, 1999; Lindberg et al., 2005b), but lower compared to a large closed landfill in Seoul, Korea, which averaged 254±224 ng m −2 h −1 (Kim et al., 2001).For the uncovered waste, our results obtained in winter 2006 (57.5-84.5 ng m −2 h −1 ) were similar to those measured in an American landfill (70±62 ng m −2 h −1 , Lindberg et al., 2005b).When compared with the mercury emission rate at local and global background sites (typically less than 30 ng m −2 h −1 ; Wang et al., 2004;Poissant and Casimir, 1998), Hg emitted from the landfill soil cover was the same or several times higher, while the uncovered waste was up to several hundreds times higher.

Working face area
The calculated Hg emission rate by the Gaussian plume model indicated the emission rate varied from one to two orders of magnitude between landfills (1.9 mg Hg h −1 at D-S landfill to 369.0 mg Hg h −1 at G-Y landfill, Table 4).This suggested Hg emissions from working face were correlated with the MSW disposal rate (Table 1), Hg content in MSW (Table 2), and the weather conditions at each landfill (Table 4).The emission factors for D-S, J-K and G-Y landfill were 0.04, 0.63, 6.81 mg Hg t −1 MSW disposed, respectively.These results were consistent with those from 7 American landfills (0.7-6.6 mg t −1 ; Lindberg and Price, 1999; Lindberg et al., 2005b).Hg emission factors indicated that 0.07‰-3.78‰Hg in MSW was released into ambient air through the working faces, with an average loss rate of 1.63‰.When combined with the emission rate and the actual area of the working face, Hg emissions from the unit area of the working face can been obtained, with a maximum of 57651 ng m −2 h −1 at G-Y landfill, which was comparable to a landfill in Florida, USA (70 000 ng m −2 h −1 ; Lindberg and Price, 1999).The results showed that the working face has the highest intensity for Hg emissions among landfill surfaces.following three reasons.First, Hg in soil pore air inside the landfill was replaced by rain water.Second, the emission pathway of landfill surface was blocked by the rainfall, more LFG was discharged through the vent pipe system.Third, the atmospheric pressure dropped on rainy days, which promoted LFG emission from the passive vent pipes.Many studies reported the flow rate of LFG from passive vent pipes was susceptible to the fluctuation of atmospheric pressure (Gebert and Groengroeft, 2006;Maurice and Lagerkvist, 2003), which generally decreased during the rainfall.We observed an obvious LFG plume from the vent pipes during rainy days as shown in Fig. 2d, whereas imperceptible emissions were found on sunny days.TGM concentrations in LFG increased several fold during the rainfall (Fig. 9), while Hg surface-air flux declined during the rainfall (Fig. 6).
Compared with TGM concentrations in LFG measured at some American landfills, our results were much lower (Lindberg and Price, 1999;Lindberg et al., 2001Lindberg et al., , 2005a;;Hawkins and Prestbo, 2004;Prestbo et al.,2003).In the latter sites, TGM concentrations was approximately at µg m −3 level, with the highest of about 12 µg m −3 .The results obtained from G-Y and J-K landfill were comparable to landfills in  Sweden (Sommar et al., 1999), Germany (Feldmann et al., 1994), South Korea (Kim and Kim, 2002) and Mexico (de la Rosa et al., 2006), ranging from several to a few of thousand of ng m −3 .The great discrepancy of TGM in LFG among different landfills may be due to mercury content in MSW and the method for LFG venting (i.e., active or passive system).

MMHg
MMHg in LFG of some vent pipes at G-Y landfill varied between 0.14 and 6.37 ng m −3 , with an average of 1.93 ng m −3 (Fig. 10).The percentage of MMHg to TGM ranged from 0.14 to 1.68%, with an average of 0.51%.The global background concentrations of MMHg in the atmosphere are generally below 10 pg m −3 (Munthe et al., 2003), thus the MMHg in LFG was about 3 orders of magnitude higher than the global background values in ambient air, showing landfill was an important MMHg emission source.MMHg concentrations of some American landfills ranged from 1-40 ng m −3 , accounting for approximately 0.1% of TGM (Lindberg et al., 2001(Lindberg et al., , 2005a).These concentrations were much higher than the data observed at G-Y landfill.

DMHg
For the same vent pipes sampled for MMHg, DMHg ranged from 2.54-19.05ng m −3 , with an average of 9.21 ng m −3 (Fig. 11).DMHg comprised 0.27 to 3.64% of TGM in the LFG, with an average of 1.79%.DMHg was also detected in LFG in the USA with concentration between 0.2 to 637 ng m −3 (Hawkins and Prestbo, 2004;Prestbo et al., 2003;Lindberg et al., 2001Lindberg et al., , 2005a)), which were much higher than that observed at G-Y landfill.DMHg is the most toxic mercury species (Nierenberg et al., 1998), and direct emissions from the landfill site could pose a serious ecological risk.It is highly recommended that LFG in China be utilized, or at least burned before it is discharged into the atmosphere.The latter method will decompose methylated Hg to elemental Hg at high temperatures, reducing their toxicity.However, from a long run, it's much better to reduce Hg in products and prevent the Hg-containing waste entering into landfills, this reduction in Hg from the source is the ultimate way to eliminate the Hg risk at landfill sites.

Summary and conclusions
Based on field experiments, Hg emission patterns from landfills were estimated (Table 5).Total Hg emissions from the five landfills in 2004 ranged from 17 to 3285 g yr −1 , with  the highest at G-Y landfill, and the lowest at X-R-J landfill.At G-Y, Hg emissions were dominated by the working face, which accounted for 98.36% of the total, followed by soil cover (1.28%), uncovered MSW (0.33%), and venting pipes (0.03%).A similar pattern was also found at J-K landfill.This confirmed that the working face was the leading source for Hg emissions from landfills, and total emissions from the vent pipes were relatively small.A rough picture for Hg emissions from all landfill sites in China can be obtained by taking into account the total MSW treated by landfill each year, total landfill surface area, total LFG generated and the meteorological data.The estimated Hg emissions from the Chinese landfill sites ranged from 500-800 kg yr −1 under different scenarios.These emission fluxes were relatively low compared to the total emissions of 552-696 tonnes Hg yr −1 from Chinese anthropogenic sources between 1995-2003(Wu et al., 2006)).However, based on the limited landfill surface, Hg emission intensity per unit area (up to 57 651 ng m −2 h −1 ) still cannot be overlooked.
In conclusion, Hg emissions from landfill sites depended on the mercury content in the substrate, were maximized at the working face, and were remarkably reduced by applying soil covering or vegetation.Hg emissions from the landfill surface were sensitive to meteorological parameters, especially solar radiation.In comparison to the vent pipe system, Hg emissions from landfill surfaces were the primary pathways.Methylated Hg species produced inside the landfill was especially important, indicated the environmental conditions (such as pH, Eh, T, O 2 level), microbial activities (such as sulfate-reducing bacteria which produce reduced sulfur compounds at landfill; Kim et al., 2005), nutrient levels, and the methyl group (CH 3− ) donors, possibly enhanced Hg methylation.However, the exact mechanisms, whether biological or chemical, are still unknown and further research is needed.

Fig. 1 .
Fig. 1.Sketch map showing the locality of five studied landfills.

Fig. 2 .
Fig. 2. Photos showing the operation landfill at G-Y (A) and closed landfill at X-R-J (B), the DFC method for surface-air flux at the soil covering area (C) and the vent pipe system for the LFG at G-Y landfill (D).

Fig. 3 .
Fig. 3.The range (A) and average (B) of total gaseous mercury in the atmosphere over different sites of the studied landfills (Sample sites: 1-22 at G-Y landfill; 23-27 at D-Z-W landfill; 28 at X-R-J landfill; 29-31 at J-K landfill and 32 at D-S landfill).

Fig. 4 .
Fig. 4. Mercury surface-air fluxes at un-covered MSW site of G-Y landfill: the contribution of high mercury content in MSW to the mercury emission rate.

Fig. 5 .
Fig. 5. Mercury surface-air fluxes at soil covering area of D-Z-W and grassed area of X-R-J landfill: the effect of vegetation on the mercury emission reduction.

Fig. 6 .Fig. 6 .
Fig. 6.Mercury surface-air fluxes at soil covering area of J-K landfill: the effect of weather 2 conditions on the mercury emission rate 3 Fig. 6.Mercury surface-air fluxes at soil covering area of J-K landfill: the effect of weather conditions on the mercury emission rate.

Fig. 7 .
Fig. 7. Correlation between average mercury flux at different sites and mercury content in the corresponding substrates.

Fig. 9 .
Fig. 9. Comparison of average TGM in the LFG before and after the rainfall event at G-Y landfill.

Fig. 10 .
Fig. 10.MMHg in the LFG of selected vent pipes at G-Y landfill.

Fig. 11 .
Fig. 11.DMHg in the LFG of selected vent pipes at G-Y landfill.

Table 1 .
Basic situation of the 5 studied MSW landfills.

Table 2 .
Statistical summaries of Hg, pH and organic matter in the MSW and cover soil at 5 studied landfills.

Table 5 .
Estimation of annual Hg emissions from the studied landfills in 2004.Landfill Emission pathways Hg emission quantities (g yr −1 ) Percentage of each pathway (%)