Meteorological parameters monitored in Kaohsiung City during the atmospheric mercury sampling periods for wet and dry seasons.
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More than half of the publishers listed alongside IntechOpen (18 out of 30) are Social Science and Humanities publishers. IntechOpen is an exception to this as a leader in not only Open Access content but Open Access content across all scientific disciplines, including Physical Sciences, Engineering and Technology, Health Sciences, Life Science, and Social Sciences and Humanities.
\\n\\nOur breakdown of titles published demonstrates this with 47% PET, 31% HS, 18% LS, and 4% SSH books published.
\\n\\n“Even though ItechOpen has shown the potential of sci-tech books using an OA approach,” other publishers “have shown little interest in OA books.”
\\n\\nAdditionally, each book published by IntechOpen contains original content and research findings.
\\n\\nWe are honored to be among such prestigious publishers and we hope to continue to spearhead that growth in our quest to promote Open Access as a true pioneer in OA book publishing.
\\n\\n\\n\\n
\\n"}]',published:!0,mainMedia:null},components:[{type:"htmlEditorComponent",content:'
Simba Information has released its Open Access Book Publishing 2020 - 2024 report and has again identified IntechOpen as the world’s largest Open Access book publisher by title count.
\n\nSimba Information is a leading provider for market intelligence and forecasts in the media and publishing industry. The report, published every year, provides an overview and financial outlook for the global professional e-book publishing market.
\n\nIntechOpen, De Gruyter, and Frontiers are the largest OA book publishers by title count, with IntechOpen coming in at first place with 5,101 OA books published, a good 1,782 titles ahead of the nearest competitor.
\n\nSince the first Open Access Book Publishing report published in 2016, IntechOpen has held the top stop each year.
\n\n\n\nMore than half of the publishers listed alongside IntechOpen (18 out of 30) are Social Science and Humanities publishers. IntechOpen is an exception to this as a leader in not only Open Access content but Open Access content across all scientific disciplines, including Physical Sciences, Engineering and Technology, Health Sciences, Life Science, and Social Sciences and Humanities.
\n\nOur breakdown of titles published demonstrates this with 47% PET, 31% HS, 18% LS, and 4% SSH books published.
\n\n“Even though ItechOpen has shown the potential of sci-tech books using an OA approach,” other publishers “have shown little interest in OA books.”
\n\nAdditionally, each book published by IntechOpen contains original content and research findings.
\n\nWe are honored to be among such prestigious publishers and we hope to continue to spearhead that growth in our quest to promote Open Access as a true pioneer in OA book publishing.
\n\n\n\n
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He was first appointed as Assistant Professor and then promoted to Associate Professor at North South University in 2011 and later on Professor. While in that post he was also the coordinator of EEE program. During 2012-2017, he was an Associate Professor at Universiti Teknologi Brunei (UTB), Brunei Darussalam (QS World University ranking 379). He received his B.Sc. degree in Electrical and Electronic Engineering from BUET (Bangladesh), his M.Sc. degree in Digital Communication from Loughborough University, UK and PhD in Wireless Communication from Newcastle University, UK. He has taught several courses in communications, electronics and signal processing at KUET, Khulna University, BRAC University, and UKM (Malaysia) during his career. He has published over 90 peer-reviewed journals and conference papers, and is the author/editor of 16 (sixteen) academic books such as Towards Cognitive IoT Networks (Springer, 2020), Communication Systems for Electrical Engineers (Springer, 2018), Spectrum Access and Management for Cognitive Radio Networks (Springer, 2016), Coding for MIMO-OFDM in Future Wireless Systems (Springer, 2015), Advances in Sensor Networks Research (Nova publisher, USA, 2014) and 10 (ten) book chapters. He has presented invited talks in Bangladesh and Malaysia and has served as a member of the program committee for more than 50 international conferences. He is on the editorial board of several international journals such as IEEE Communications Magazine, IEEE, USA, IET Wireless Sensor Systems (IET-WSS), and so on. Dr. Matin is a member of the IEEE, IEEE Communications Society (IEEE ComSoc), and several other international organizations. 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The United States had already listed 189 hazardous air pollutants (HAPs) for control in Title III of the Clean Air Act Amendments (CAAA) of 1990 [1, 2]. Among them, 11 are the toxic heavy metals As, Be, Cd, Co, Cr, Hg, Ni, Mn, Pb, Sb, and Se. In 2005, the U.S.A. Environmental Protection Agency (USEPA) became the first agency to amend the Maximum Achievable Control Technology (MACT) standards into growth control quotas for mercury emissions in order to increase the original goal of a 30% reduction in mercury emissions to 70% by 2018 [2]. This was also done to encourage countries around the world to revise their own reduction goals for mercury emissions. Till now, the pollutants of mercury still cannot be completely removed by using chemical method, and continue to endanger the health of humans and other organisms.
\nMercury (Hg) is a persistent, toxic, and bio-accumulative heavy metal, and is currently regulated by the USEPA and the United Nations Environment Programme (UNEP) [3, 4, 5, 6, 7, 8, 9, 10], and atmospheric mercury has been claimed by UNEP as another global environmental issue followed greenhouse gases (GHGs) [1, 2]. Because of its unique physicochemical properties and potential for long-range transportation via the atmospheric dispersion pathway, it could be deposited worldwide [4, 5]. Due to its characteristics of persistence and bioaccumulation through food chain, mercury could damage the brain and nervous systems in human body. Thus, many countries are becoming increasingly concerned about atmospheric mercury pollution recently.
\nAccordingly, this chapter aimed to investigate the tempospatial variation and partition of atmospheric mercury at an industrial metropolitan area, and to explore the long-range transportation path at the marine boundary layer (MBL), and to compare the atmospheric mercury concentration level with other major cities all over the world.
\nThe mercury emission was mainly generated from combustion sources and metallurgy smelting processes in Taiwan. Among them, the combustion sources include coal-fired boilers, municipal and industrial waste incinerators, petrochemical refineries, cremators, etc., while the metallurgy smelting processes include integrated steel plants, electric arc furnace plants, secondary metal smelters, etc. The economic development of metro Kaohsiung mainly relied on heavy industries located adjacent to the metropolitan area. Coupled with heavy traffics, it causes poor ambient air quality of Kao-Ping Air Quality Zone in Taiwan. Moreover, metro Kaohsiung is the largest and most intensive industrial city in Taiwan, accounting for about 70% of the stationary emission sources, where has two large-scale coal-fired power plants, an integrated steel plant, several metallurgy smelters, and three petrochemical refineries. It would cause quite high emissions of mercury in metro Kaohsiung, Taiwan. In this chapter, the field measurement of atmospheric mercury was designated to investigate the tempospatial variation and partition of gaseous and particulate mercury during the wet and dry seasons, and further correlated atmospheric mercury with meteorological parameters and criteria air pollutants measured in Kaohsiung City, located at the coastal region of southern Taiwan.
\nField measurement of atmospheric mercury speciation and concentration was conducted at a coastal background site and six sensitivity sites in Kaohsiung City, including Nan-zhi (22°44′00″ North latitude, 120°19′41″ East longitude, S1), Ren-wu (22°41′20″ North latitude, 120°19′57″ East longitude, S2), Zuoying (22°40′29″ North latitude, 120°17′34″ East longitude, S3), Cia-jin (23°37′56″ North latitude, 120°17′16″ East longitude, S4), Cianjhen (22°36′18″ North latitude, 120°18′30″ East longitude, S5), and Hsiao-kang (22°33′57″ North latitude, 120°20′15″ East longitude, S6). The coastal background site along coastline far away from the emission sources is located at the campus of National Sun Yat-Sen University (22°37′38″ North latitude, 120°16′01″ East longitude). The location of the coastal background site and six sensitivity sites for sampling total gaseous mercury (TGM) and particulate mercury (Hgp) in Kaohsiung City is shown in Figure 1.
\nThe location of six sensitivity sites and the coastal background site for TGM and Hgp sampling in Kaohsiung City.
These sensitivity sites were mainly located at the ambient air quality monitoring stations in Kaohsiung City. Among them, sites S5–S6 were nearby a steel industrial complex in southern Kaohsiung, sites S1 and S2 were close to a petrochemical industrial complex in northern Kaohsiung. Active sampling of TGM and Hgp were conducted consecutively for 24 h at each site in the wet and dry seasons from June to December of 2010 in Kaohsiung City. Among them, wet season started from June to September, while dry season started from October to December. This study intended to investigate the seasonal variation, the spatial distribution, and the partition of TGM and Hgp at a coastal background site and six sensitivity sites in an industrial city.
\nFigure 2 illustrates the seasonal variation of TGM and Hgp concentrations at the northern and southern Kaohsiung as well as the coastal background site during the wet and dry seasons. The field measured meteorological data are listed in Table 1, and the field measurement data with mean, standard deviation (SD), and partition of TGM and Hgp are summarized in Table 2.
\nTempospatial variation of TGM and Hgp concentrations during the wet and dry seasons in the northern and southern Kaohsiung sites and the coastal site.
Seasons | \nTemp. (°C) | \nRH (%) | \nRainfall (mm) | \nRainy days | \nUVB (UVI) | \nWS (m/s) | \nWD | \n
---|---|---|---|---|---|---|---|
Wet | \n28.9 ± 0.7 | \n78.3 ± 1.7 | \n427.9 ± 305.1 | \n50 | \n3.4 ± 0.5 | \n2.2 ± 0.2 | \nSW, NE | \n
Dry | \n23.7 ± 3.4 | \n73.3 ± 3.1 | \n63.0 ± 97.6 | \n12 | \n2.1 ± 0.3 | \n1.9 ± 0.1 | \nNW, NE | \n
Meteorological parameters monitored in Kaohsiung City during the atmospheric mercury sampling periods for wet and dry seasons.
Temp.: ambient temperature; RH: relative humidity; Rainy days: days with rainfall ≧ 0.01 mm; WS: wind speed; WD: wind direction; UVB: ultra-violet radiation.
Seasons | \nTypes of Hg | \nNorthern Kaohsiung sites | \nSouthern Kaohsiung sites | \nCoastal background site | \n||||
---|---|---|---|---|---|---|---|---|
S1 | \nS2 | \nS3 | \nS4 | \nS5 | \nS6 | \nSb | \n||
Wet season (n = 12) | \nTGM (ng/m3) | \n5.72 | \n6.37 | \n5.20 | \n4.04 | \n7.42 | \n8.60 | \n2.38 | \n
Hgp (ng/m3) | \n0.20 | \n0.24 | \n0.06 | \n0.05 | \n0.35 | \n0.50 | \n0.02 | \n|
Mean ± SD (ng/m3) | \nTGM | \n6.23 ± 1.62 | \n\n | |||||
Hgp | \n0.23 ± 0.17 | \n\n | ||||||
Dry season (n = 12) | \nTGM (ng/m3) | \n5.95 | \n8.06 | \n6.59 | \n5.03 | \n7.50 | \n9.41 | \n2.44 | \n
Hgp (ng/m3) | \n0.23 | \n0.39 | \n0.12 | \n0.08 | \n0.59 | \n0.68 | \n0.03 | \n|
Mean ± SD (ng/m3) | \nTGM | \n7.09 ± 1.57 | \n\n | |||||
Hgp | \n0.35 ± 0.25 | \n\n | ||||||
Mean ± SD of atmospheric mercury (ng/m3) | \nTGM | \n6.66 ± 1.42 | \n2.41 ± 0.04 | \n|||||
Hgp | \n0.29 ± 0.21 | \n0.03 ± 0.01 | \n
Seasonal variation of TGM and Hgp concentrations at six sensitivity sites and the coastal background site in Kaohsiung City.
S1: Nan-zhi site; S2: Ren-wu site; S3: Zuo-ying site; S4: Cia-jin site; S5: Cian-jhen site; S6: Hsiao-kang site; Sb: coastal site.
During the wet season, ambient temperature, relative humidity, wind speed, and UVB were 28.9 ± 0.7°C, 78.3 ± 1.7%, 2.2 ± 0.2 m/s, and 3.4 ± 0.5 UVI, respectively, which were mostly higher than those during the dry season (Table 2). The prevailing wind blew from southwest and northeast during the dry season, and the prevailing wind blew from northwest and northeast during the wet season were reported. It was mainly attributed to the fact that sea-land breezes blew frequently in Kaohsiung City. However, the wet season is the heat convection season (i.e., hurricane season), in which the rainfall was about 427.9 ± 305.1 mm and the rainy days (rainfall ≧ 0.01 mm) were about 50 days. The dry season commonly blew the northeast monsoon, in which the rainfall and rainy days were much less than those during the wet season. Therefore, the aforementioned meteorological condition is considered as the differential between the dry and wet seasons in Kaohsiung City.
\nField measurement data showed that the concentrations of TGM and Hgp were 6.23 ± 1.62 ng/m3 and 0.23 ± 0.17 ng/m3, respectively, during the wet season, while the concentrations of TGM and Hgp were 7.09 ± 1.57 ng/m3 and 0.35 ± 0.25 ng/m3, respectively, during the dry season in Kaohsiung City. The concentrations of atmospheric mercury during the dry season was obviously higher than those during the wet season. It showed that meteorological condition and atmospheric dispersion played a critical role on the seasonal variation of atmospheric mercury concentration. However, the seasonal concentrations of TGM and Hgp did not vary much at the coastal background site, thus the seasonal variation has insignificant influences on regions where atmospheric mercury concentrations were high.
\nThe atmospheric mercury concentrations at southern Kaohsiung were mostly higher than those at northern Kaohsiung during the wet and dry seasons, and their average concentrations were respectively 1.12 and 1.79 times of those at northern Kaohsiung. As a whole, the average concentrations of TGM and Hgp in Kaohsiung City were about 2.94 and 11.7 times, respectively, higher than those at the coastal background site during the wet season, and were about 2.6 and 11.5 times, respectively, during the dry season. Overall, the average TGM and Hgp concentrations were 6.66 ± 1.42 and 0.29 ± 0.21 ng/m3, respectively, in Kaohsiung City. The TGM concentration in Kaohsiung City was about 4.2 times and 2.8 times higher than the background TGM concentration of the North Hemisphere (1.6 ng/m3) and at the coastal background site in Kaohsiung City (2.4 ng/m3), respectively. It showed that Kaohsiung City as a heavy industrial city was highly polluted by atmospheric mercury.
\nFigure 3 illustrates the partition of Hgp during the wet and dry seasons in Kaohsiung City. The results showed that TGM was the main mercury species, accounting for 94.56–99.59% of atmospheric mercury during the wet season, and 92.71–99.37% of atmospheric mercury during the dry season. Furthermore, Hgp concentration had a tendency to increase with the distance from the emission sources. The maximum partition of Hgp were up to 20–40% of total atmospheric mercury (TAM) in the ambient air [11, 12, 13]. The partition of Hgp was generally lower than 1% of TAM in the rural areas, about 1–3% in the metropolitan areas, and beyond 5% in the industrial areas [14, 15, 16]. Figure 1 shows that site Sb is in the rural area, sites S3 and S4 are in the metropolitan areas, and other four sites are in the industrial areas in Kaohsiung City. Moreover, the partition of Hgp during the dry season was higher than that during the wet season. It implied that the amount of rainfall, the number of rainy days, and relative humidity might correlate to the partition of Hgp.
\nPartition of Hgp during the wet and dry seasons at six sensitivity sites (S1–S6) and the coastal site (Sb).
The concentrations of TGM and Hgp measured at each site are summarized in Table 2. As far as the spatial distribution of atmospheric mercury during the wet and dry seasons in Kaohsiung City was concerned, the atmospheric mercury concentrations (TGM of 6.37 ng/m3; Hgp of 0.24 ng/m3) at site S2 was the highest and followed by sites S1 and S3 in northern Kaohsiung, while those (TGM of 8.60 ng/m3; Hgp of 0.50 ng/m3) at site S6 was the highest and followed by sites S5 and S4 in southern Kaohsiung. During the dry season, the atmospheric mercury concentrations (TGM of 8.06 ng/m3; Hgp of 0.39 ng/m3) at site S2 was also the highest and followed by sites S3 and S1 in northern Kaohsiung, while those (TGM of 9.41 ng/m3; Hgp of 0.50 ng/m3) at site S6 was the highest and followed by sites S5 and S4 in the southern Kaohsiung.
\nThe mapping software (SURFER) was further used for plotting the concentration contour of atmospheric mercury in Kaohsiung City. This software uses the grid difference as the calculation basis, and interpolarates the data of atmospheric mercury concentration obtained from each sampling site into the map of Kaohsiung City. This study aimed to explore the spatial distribution of atmospheric mercury concentration in Kaohsiung City. As shown inFigure 4, the atmospheric mercury concentrations of northern and southern Kaohsiung were obviously affected by the mercury emission sources. Two major high mercury concentration regions concurred with the petrochemical industrial district in northern Kaohsiung and the steel manufacturing complex in southern Kaohsiung. The main emission sources in northern Kaohsiung were petrochemical manufacturing complex, municipal and industrial waste incinerators, cremators, etc. In southern Kaohsiung, mercury was mainly emitted from steel smelters, electric arc furnaces, cement plants, petroleum refineries, coal-fired power plants, municipal waste incinerators, etc. Consequently, the atmospheric mercury concentration in southern Kaohsiung was higher than that in northern Kaohsiung, which was attributed to higher mercury emission in southern Kaohsiung than that in northern Kaohsiung. As a whole, the ambient air quality of Kaohsiung City was highly affected by the heavily densed industries, thus the concentration of atmospheric mercury in the metropolitan area was much higher than the background level during the wet and dry seasons. Both numbers of the emission sources and the consumption of fossil fuels in southern Kaohsiung were higher than those in northern Kaohsiung.
\nSpatial distribution of atmospheric mercury in Kaohsiung City. (A) TGM concentration during the dry season, (B) TGM concentration during the wet season, (C) Hgp concentration during the dry season, and (D) Hgp concentration during the wet season.
Figure 5 compares the concentrations of TGM and Hgp at different mercury emission sources in southern Taiwan [12, 13]. The highest TGM concentrations were observed at a steel plant, which was approximately 2.6 times higher than those at Tainan Scientific Complex, even up to 15 times for Hgp concentration. Similarly, other stationary sources were also higher than those observed at Tainan Scientific Complex for both TGM and Hgp concentrations. Figure 6 illustrates the partition of TGM and Hgp for various mercury emission sources in southern Kaohsiung. The partition of Hgp at Tainan Scientific Complex was only 1.7%, which was similar to those observed at most urban areas (1–3%), and was much lower than other stationary sources (e.g., coal fired power plants, waste incinerators, and steel plants). Their partition of Hgp ranged from 7.3 to 9.5%, which contributed much more Hgp to the ambient air [6, 15, 17, 18]. Overall speaking, TGM was the dominant species of atmospheric mercury in southern Taiwan.
\nComparison of TGM and Hgp concentrations with various mercury emission sources in southern Kaohsiung (A: the semi-conductor complex, B: the petroleum refinery, C: the steel plant, D: the coal-fired power plant, E: the electric arc furnace, F: the municipal waste incinerator).
Location of the TGM monitoring site (★) at the Penghu Islands.
Penghu Islands are located at the middle of the Taiwan Strait between the southeastern China and Taiwan. The islands are dominated by subtropical weather and mainly influenced by East Asian monsoons. Previous studies reported that, during the seasons of spring and winter, the biomass and fossil fuel burning frequently occurring in Southeast Asia and Southwest China increase the levels of atmospheric mercury in Taiwan owning to long-range transportation [11, 18]. Since there are no significant mercury sources at the Penghu Islands, it can be treated as the background site in the region.
\nIn this chapter, a one-year field monitoring protocol was conducted to investigate the seasonal and daily variations of TAM concentration at the Penghu Islands, as the correlation of TGM concentration with the meteorological parameters and several criteria air pollutants being further examined and discussed. More importantly, a backward trajectory simulation model was further applied to explore the transportation of TGM to the Penghu Islands with respect to the transportation routes for those observed at 500 m above the sea level during the monitoring seasons. While the TGM has been a continuing issue of great concern worldwide, the results of this study would help development more effective management strategies to control the adverse influences associated with the effect of TGM on the environment and human health as well in the Penghu Islands and the areas possibly affected by the long-range transportation of TGM.
\nAccording to the meteorological data obtained from 1985 to 2011, dry season with the rainfall of about 800 mm started from April to September. Table 3 summarizes the meteorological data measured at the Penghu Islands during the TGM monitoring seasons, indicating that the prevailing winds were blown from the east, northeast, northwest, and south, with the ambient temperatures of 13.4–31.2°C, the relative humidity of 53.2–91.3%, and the wind speeds of 0.5–9.7 m/s. The TGM monitoring site was located on the roof of a four-floor building, which was approximately 12 m above the ground and 50 m and 500 m far from the coastline and the major roads, respectively. Overall, the TGM monitoring site was located at the marine boundary layer (MBL), reducing the possible interferences resulted by Hg emission sources. The atmospheric TGM was continuously monitored at the Hsiaomen site (23°38′471″ North latitude, 119°30′316″ East longitude) located at the northwestern coastline of the Penghu Islands (see Figure 7) for 15 continuous days in each season from March of 2011 to February of 2012.
\nSeasons | \nn | \nTemp. (°C) | \nRH (%) | \nWS (m/s) | \nWD | \n
---|---|---|---|---|---|
Spring | \n360 | \n19.9 ± 2.1 | \n72.5 ± 7.5 | \n4.5 ± 2.2 | \nENE | \n
Summer | \n360 | \n29.8 ± 0.6 | \n79.3 ± 2.7 | \n2.2 ± 0.6 | \nS | \n
Fall | \n360 | \n24.8 ± 0.6 | \n75.1 ± 4.9 | \n6.1 ± 1.5 | \nENE | \n
Winter | \n360 | \n16.1 ± 1.5 | \n82.2 ± 4.7 | \n6.2 ± 1.3 | \nENE | \n
Mean ± SD | \n22.7 ± 5.9 | \n77.3 ± 4.3 | \n4.8 ± 1.9 | \n— | \n|
Max | \n33.2 | \n91.3 | \n9.7 | \n— | \n|
Min | \n13.4 | \n53.2 | \n0.5 | \n— | \n
Meteorological data recorded at the Penghu Islands during the TGM sampling periods.
Temp.: ambient temperature; RH: relative humidity; WS: wind speed; WD: wind direction; SD: standard deviation.
Variation of TGM concentration during four monitoring seasons at the Penghu Islands.
This chapter collected the meteorological data and the criteria air pollutant concentrations from the Makung Air Quality Station during the TGM monitoring periods, and discussed the correlation coefficients derived from the TGM concentrations with the meteorological parameters and the criteria air pollutants. Among them, the correlation coefficient (R) was used to describe the correlation between TGM concentration, meteorological data, and criteria air pollutant concentrations. The meteorological parameters included the ambient temperature (Temp.), relative humidity (RH), wind speed (WS), and wind direction (WD), while the criteria air pollutants of interest included SO2, NOX, CO, O3, PM10, and PM2.5. Four seasons are defined as March to May (spring), June to August (summer), September to November (fall), and December to February (winter), respectively. 72-h backward trajectories were simulated by using a NOAA HYSPLIT model with the National Centers for Environmental Prediction’s Global Data Assimilation System (NCEP-GDAS) meteorological dataset used as the model input in this study. All backward trajectories started with an arrival height of 500 m above the sea level. By using the NOAA-HYSPLIT model, the dates of the highest TGM concentration at the Penghu Islands in different seasons were determined and the transportation routes of air masses toward the Penghu Islands during the monitoring periods were then simulated.
\nThis information was further applied to examine the possible transportation routes conveying TGM from the upwind sources to the Penghu Islands. Additionally, we used the local fire map to identify the possible sources of TGM drawn by the FIRMS web (
Table 4 summarizes the average and standard deviation of TGM and criteria air pollutant concentrations measured at the Penghu Islands. The highest concentrations of PM10 and PM2.5 observed in spring were 54.60 ± 15.56 and 31.07 ± 10.59 μg/m3, respectively, while the lowest concentrations of PM10 and PM2.5 occurred in summer were 36.63 ± 8.69 and 17.70 ± 6.48 μg/m3, respectively. Spring and winter are two major seasons frequently blowing Asian dusts from the northern China to the Penghu Islands [19], which significantly increased the concentrations of PM10. Overall speaking, the concentrations of SO2 and NOx measured at the Penghu Islands were much lower than other cities in Taiwan, indicating no significant local sources in the Penghu Islands. Moreover, O3 concentration was relatively higher compared to other criteria air pollutants at the Penghu Islands.
\nSeasons | \nn | \nTGM (ng/m3) | \nPM10 (μg/m3) | \nPM2.5 (μg/m3) | \nSO2 (ppb) | \nNOx (ppb) | \nO3 (ppb) | \nCO (ppm) | \n
---|---|---|---|---|---|---|---|---|
Spring | \n360 | \n4.31 ± 1.87 | \n54.60 ± 15.56 | \n31.07 ± 10.59 | \n1.70 ± 0.72 | \n5.61 ± 0.08 | \n61.58 ± 5.00 | \n0.27 ± 0.06 | \n
Summer | \n360 | \n1.81 ± 0.15 | \n36.63 ± 8.69 | \n17.70 ± 6.48 | \n2.21 ± 0.63 | \n6.15 ± 1.23 | \n33.17 ± 7.70 | \n0.15 ± 0.03 | \n
Fall | \n360 | \n3.03 ± 0.40 | \n45.24 ± 18.49 | \n21.40 ± 6.08 | \n1.68 ± 0.52 | \n3.48 ± 0.46 | \n51.86 ± 5.99 | \n0.18 ± 0.04 | \n
Winter | \n360 | \n3.51 ± 0.67 | \n48.65 ± 9.79 | \n23.81 ± 12.18 | \n2.00 ± 0.08 | \n5.48 ± 1.63 | \n38.59 ± 5.87 | \n0.28 ± 0.08 | \n
P-value | \n3.32 × 10−55 | \n6.94 × 10−3 | \n1.93 × 10−3 | \n1.91 × 10−1 | \n1.36 × 10−7 | \n4.00 × 10−18 | \n3.04 × 10−10 | \n|
Mean ± SD | \n3.17 ± 1.06 | \n46.28 ± 7.51 | \n23.50 ± 5.64 | \n1.90 ± 0.25 | \n5.18 ± 1.17 | \n46.30 ± 12.86 | \n0.22 ± 0.06 | \n|
Range | \n1.17–8.63 | \n17.18–77.68 | \n1.32–63.07 | \n1.02–8.63 | \n2.43–14.09 | \n11.35–88.08 | \n0.16–0.57 | \n
Continuous monitoring data of TGM concentrations and their mean, standard deviation (SD), range, and the concentrations of criteria air pollutants.
The concentrations observed among the seasons were significantly different by the analysis of ANOVA (Analysis of Variance) at the confidence level of 95%.
Field monitoring of TGM at the Penghu Islands showed that the average concentration of TGM was 3.17 ± 1.06 ng/m3 with the range of 1.17–8.63 ng/m3. The concentration of TGM in four monitoring seasons were ordered as: spring (4.34 ± 1.87 ng/m3) > winter (3.51 ± 0.67 ng/m3) > fall (3.03 ± 0.40 ng/m3) > summer (1.81 ± 0.15 ng/m3). Moreover, the average TGM concentration (3.17 ng/m3) at the Penghu Islands was approximately two times of the background TGM concentration of North Hemisphere (1.6 ng/m3). The Penghu Islands are likely to be influenced by long-range transportation of TGM in spring and winter, causing higher TGM concentrations in spring and winter than those in summer and fall. The lowest seasonal average TGM concentration of 1.81 ± 0.15 ng/m3 was observed in summer, which was slightly close to the background TGM concentration of North Hemisphere, suggesting that air masses blown from the South China Sea were relatively clean.
\nFigure 8 illustrates the variation of TGM concentrations during the four monitoring seasons at the Penghu Islands. The daily concentration of TGM in spring increased significantly from April 2nd to the peak TGM concentration (8.63 ng/m3) observed on April 10th. The lowest TGM concentration ranging from 1.50 to 2.70 ng/m3 was compatibly observed in summer. Moreover, the concentrations of criteria air pollutants were also the lowest in summer compared to other seasons (Table 4), suggesting that the ambient air quality in summer was relatively better than other seasons at the Penghu Islands, and the TGM concentrations was 1.7–2.8 times lower than other seasons.
\nHourly variation of TGM concentrations during four monitoring seasons at the Penghu Islands.
Unlike other seasons, the TGM concentrations fluctuated frequently in fall and winter. In winter, it increased rapidly from 3.74 to 5.08 ng/m3 on December 28th, and then decreased to 3.38 ng/m3 on December 30th. The highest peak concentration (4.35 ng/m3) occurred on December 29th and then leveled off on January 5th. Long-term TGM monitoring at Mt. Lulin background air quality monitoring station showed that Taiwan was highly influenced by atmospheric mercury and gaseous pollutants from China and Southeast Asia in spring and winter [11]. Backward trajectory simulation results indicated that the atmospheric mercury detected in Taiwan significantly increased due to frequent biomass burning originated from the Southeast Asia and industrial emissions from the North China in spring and winter. It suggested that the TGM concentrations at the Penghu Islands might be also influenced by the atmospheric mercury emitting and transporting from these regions. While, air masses blown from the Pacific Ocean had relatively low contribution to the TGM concentration.
\nFigure 9 illustrates the hourly variation of TGM concentrations during four seasons at the Penghu Islands. It showed that the variation of hourly TGM concentrations were relatively steady in summer. The TGM concentration increased from 6:00 am, gradually reached its concentration peak at 11:00 am, and then decreased after noontime. The increase of TGM concentration resulted probably from the following two processes: (a) UV radiation could temporally transform Hg+, Hg2+, and Hgp to volatile Hg0, and subsequently entered to the atmosphere [6]); (b) the downward mixing from the enhanced TGM aloft may increase the levels of TGM concentration as the destroying of nocturnal inversion layer [20]. Except for the morning time, the TGM concentration was less variable through the whole day. The average TGM concentrations in the daytime (6:00 am–6:00 pm) were typically higher than those at nighttime (6:00 pm–next 6:00 am). These findings are attributed to the effects from the variation of the height of atmospheric boundary layer. Additionally, it might be influenced by local anthropogenic activities, such as open burning, mobile sources of fishing boats, vehicles, etc.
\nFrequency distribution of TGM concentration covering the four monitoring seasons.
The magnitude of hourly variation (differentia between the maximum and the minimum TGM concentrations) was lower in summer and winter with the relative percent differences (RPD) of 61.9 and 88.1%, respectively, and higher in spring and fall with the RPD of 130.7 and 107.8%, respectively. However, previous studies showed that the monitored TGM concentration varies at low altitudes in the typical rural areas [11, 21, 22]. This study revealed that the concentration of TGM monitored at the Penghu Islands was mainly influenced by both local biomass burning and long-range transportation.
\nFigure 10 illustrates the frequency distribution of TGM detected during four monitoring seasons. It showed that the TGM concentrations followed a lognormal distribution pattern in the range of 2.0–4.5 and 6.0–7.5 ng/m3, accounting for approximately 80.0% of total frequency in spring. The values ranging from 1.5 to 2.5 ng/m3 accounted for approximately 89.4% of total frequency in summer. Similarly, the values in the range of 2.0–4.5 and 3.5–5.5 ng/m3 accounted for 93.3 and 61.7% of total frequency in fall and winter, respectively. However, the episodes with high TGM concentrations (>9.0 ng/m3) were frequently observed in spring, fall, and winter. It is worth noting that the frequency distribution of TGM appeared to follow two different trends, as shown in Figure 10. The frequency distributions of TGM levels in summer and fall were unimodal distribution, while those in spring and winter followed a multi-modal distribution.
\nPollution rose of TGM in four seasons at the Penghu Islands.
Figure 11 illustrates the pollution rose of TGM for four monitoring seasons at the Penghu Islands. High TGM concentrations were observed mainly in the wind directions of 0–90° and followed by 270–360° in spring, and 60–120° in fall and winter. There were no significant mercury emission sources at the northwestern or southwestern Penghu Islands, suggesting that the high concentrations of TGM were transported remotely from China. An increasing number of studies have shown that biomass burning, industrial combustion, and ocean evaporation are three major emission sources of TGM in the regional scale [10, 23, 24, 25]. Figure 12 illustrates the backward trajectories and the TGM concentration percentage of air masses toward the Penghu Islands in four seasons.
\nBackward trajectories of air masses and TGM concentration percentage for different pathways during the monitoring seasons at the Penghu Islands.
The percentage of the TGM concentration attributed from long-range transportation at the Penghu Islands.
In spring, the TGM concentrations for air masses transported conveying from routes (1) and (2) ranged from 3.55–7.12 ng/m3, accounting for approximately 89% of TGM data, which was possibly dominated by those transported from local stationary sources and open burning with air masses toward the Penghu Islands. Thus, the southern China were another possible sources contributing to the TGM levels at the Penghu Islands.
\nIn summer, the TGM concentrations for air masses conveying from routes (4) and (5) ranged from 1.48–2.39 ng/m3, accounting for approximately 97% of TGM data, which were transported from the South and the East China Sea to the Penghu Islands, dramatically increasing the TGM concentrations at the Penghu Islands during the summer monitoring period. Consequently, the TGM concentrations in summer were relatively lower than other seasons, and the average concentration of TGM was close to the background TGM concentration of North Hemisphere (approximately 1.6 ng/m3).
\nIn fall, the TGM concentrations for the air masses conveying from routes (7) and (8) ranged from 3.02 to 3.73 ng/m3, accounting for approximately 96% of TGM data, which seemed to be mainly transported from the northern China, Korea, and Japan with air masses toward the Penghu Islands, resulting in approximately 1.67 times higher TGM concentration in fall than those in summer.
\nIn winter, the TGM concentrations for the air masses conveying from routes (11) and (12) ranged from 3.67 to 3.84 ng/m3, accounting for approximately 85% of TGM data. The potential sources conveying the TGM toward the Penghu Islands in winter were from the southern Asia, northern China, and Mongolia, increasing the TGM concentrations up to 1.94 times of those in summer.
\nIn this study, the minimum TGM value of 1.5 ng/m3 in the summer monitoring period, can be treated as the background concentration of TGM at the Penghu Islands. Therefore, the TGM concentration attributed from long-range transportation can be obtained by Eq. (1), and illustrated in Figure 13:
\nFire maps air mass during four monitoring seasons in the East Asia in (a) spring, (b) summer, (c) fall, and (d) winter.
where contribution percentage is the TGM concentration attributed from cross-boundary transportation (%); VMH is the monitored TGM concentrations (ng/m3); BC is the background concentration of Hg (i.e., 1.5 ng/m3). The contribution percentage of TGM concentration attributed from cross-boundary transportation at the Penghu Islands were further compared. The percentages of cross-boundary transportation were ordered as: spring (59.6 ± 15.0%) > winter (55.9 ± 7.6%) > fall (49.7 ± 6.2%) > summer (16.8 ± 6.9%). In addition, the maximum percentage of 79.1% was observed in spring. It showed that the TGM concentrations in spring at the Penghu Islands were highly influential, which were mainly affected by the long-range transportation.
\nHigh humidity and rainfall frequency at the Penghu Islands could also explain relatively lower TGM concentrations measured in winter. Previous study reported that the occurrence of biomass burning such as forest fires from February to April in the Southeast Asia and the Indochina Peninsula emitting a large amount of mercury-containing pollutants to the atmosphere [26]. Figure 14 illustrates the fire maps obtained from the FIRMS web fire maps and air mass transportation routes in four seasons. These fire maps explained why high TGM concentration occurred in spring at the Penghu Islands. During the monitoring periods, the hot spot of fires occurred densely in spring than other seasons could emit a large amount of TGM to the atmosphere, and then transported toward the Penghu Islands.
\nFThe concentration map of TGM at islands and seas in East Asia. (a) An-Myun Island (Nguyen et al., 2007); (b) Yellow Sea (Ci et al., 2011); (c) Jeju Island (Nguyen et al., 2010); (d) Okinawa Island (Chand et al., 2008); (e) Penghu Islands (This study); and (f) South China Sea (Fu et al., 2010).
As illustrated in Figure 15, the TGM concentrations monitored at the Penghu Islands were compared to those observed at islands and seas in East Asia. The TGM concentrations in the ambient air were ordered as: An-Myun (4.61 ± 2.21 ng/m3) > Jeju (3.85 ± 1.68 ng/m3) > Penghu (3.17 ± 1.06 ng/m3) > Yellow Sea (2.61 ± 0.50 ng/m3) > South China Sea (2.32 ± 2.62 ng/m3) > Okinawa Islands (2.04 ± 0.38 ng/m3) [10, 22, 25, 26, 27]. The results showed that the TGM concentrations decreased continuously from the northern islands to the southern islands. The TGM concentrations observed at the offshore islands were generally lower than those close to the continent or main island (except Yellow Sea). It further explained why the high levels of TGM concentration at the offshore islands were much easily influenced by Hg emission sources from the anthropogenic sources in the continent or main island.
\nThe concentration map of TGM at islands and seas in East Asia. (a) An-Myun Island; (b) Yellow Sea; (c) Jeju Island; (d) Okinawa Island; (e) Penghu Islands; and (f) South China Sea.
In summary, in addition to the local sources and open burning, the concentration of TGM at the Penghu Islands was mainly influenced by the long-range transportation of air masses, as the prevailing wind direction and air mass transportation routes potentially playing the critical roles on the variation of TGM concentration in the atmosphere.
\nThis chapter investigated the atmospheric mercury by using the modified sampling and analytical methods from two cases of small-scale regions to large-scale regions, respectively, which further investigated the tempospatial variation of atmospheric mercury, gas-particulate partition, transportation routes of mercury, and comparison of mercury concentration in urban areas and stationary sources. According to the results of this field study, several major conclusions are summarized as follows.
\nThe tempospatial variation and the partition of TGM and Hgp in Kaohsiung City were 6.66 ± 1.42 ng/m3 and 0.29 ± 0.21 ng/m3, respectively. The TGM concentration was approximately 4.1 times higher than the background concentration of 1.6 ng/m3 in North Hemisphere. Two high mercury concentration regions in Kaohsiung City concurred with the petrochemical complex in Northern Kaohsiung, and the steel manufacturing complex in Southern Kaohsiung.
\nThe TGM and Hgp concentrations in Kaohsiung City were generally higher than those of other Taiwanese cities during the wet and dry seasons. The concentrations of TGM measured in the cities of Taiwan were generally higher than Tokyo and Seoul, however, lower than other cities in China. The burning of coal for space heating in wintertime makes China the main mercury emission source in the world. Moreover, Japan, Korea, and Taiwan are under the leeward of China, which are also in the major atmospheric mercury transportation routes. When the northeastern monsoon prevails, it resulted in the increase of TGM and Hgp concentrations in the cities of East Asia.
\nTGM concentration monitored at the Penghu Islands was 3.17 ± 1.06 ng/m3 with the range of 1.17–8.63 ng/m3, and were ordered as: spring > winter > fall > summer. Summer is the only season close to the background TGM concentration of Northern Hemisphere at the Penghu Islands. While the hourly variation, TGM concentration typically increased in the morning (8:00 am–1:00 pm), reached its peak concentration, and then decreased in the late afternoon (after 2:00 pm).
\nAir masses transported from the southern and northern China, the southern Asia, Korea, Japan, and Mongolia might affect the Hg levels at the Penghu Islands during the monitoring seasons. The concentrations of TGM might be influenced by the mercury-polluted air masses to be transported remotely from areas or local stationary combustion and mobile sources. While air masses transported toward the Penghu Islands was dominated by that transported from South China Sea in summer, the TGM concentration levels at the Penghu Islands appeared to be lower than other seasons.
\nHigh TGM concentration observed at the Penghu Islands in spring might be attributed to the following three reasons: (a) local emissions from field open burning, local stationary combustion, and mobile sources; (b) long-range transportation from biomass burning in Southeast Asia or neighboring Chinese coastal cities, and (c) long-range transportation through Asian dusts from North China.
\nThe authors gratefully acknowledge the financial support and kind assistance from National Sun Yat-Sen University of Air Pollution Control Laboratory (APCL). The authors would like to express their sincere appreciation for its financial support to accomplish this study.
\nThe intensive use of organochlorine pesticides (OCPs) during the last decades around the world and the inadequate management of the wastes generated during the production of these compounds represents a huge environmental problem. That is the case of lindane production, the gamma isomer of hexachlorocyclohexane (γ-HCH), whose production during the last century has generated large amounts of solid wastes, consisting of a mixture of other HCH isomers, that has caused hot points of soil and groundwater contamination [1].
\nLindane was synthesized for the first time in 1825 by Michael Faraday [2] and deeply used as a broad-spectrum organochlorine insecticide since the 1940s [3, 4]. Among the eight isomers of HCH, lindane is the only one with insecticidal properties. Unfortunately, the lindane production, schematically summarized in Figure 1, is an inefficient process, generating large volumes of the other HCH isomers (mainly α-, β- and δ-HCH). The mixture of HCH isomers obtained in the chlorination of benzene is called technical-HCH, and it was usually subjected to a purification process to separate the γ-HCH isomer. After this step, about 10 kg of HCH wastes were obtained per kg of purified lindane. The solid HCH wastes (consisting of a white powder of HCH isomers) were inappropriately dumped during decades in the production sites nearby, resulting in environmental contamination with global dimension [5, 6, 7, 8, 9, 10, 11, 12, 13, 14].
\nScheme of lindane production and purification processes.
It is estimated that approximately 450,000 tons of lindane were used worldwide between 1950 and 2000. Approximately 63% of the lindane produced was consumed in Europe, 17% in Asia, and about 4.2% in the United States, resulting in the ubiquitous presence of HCH wastes, as is shown in Figure 2.
\nLocation of sites polluted with HCH wastes around the world, modified from [10].
Due to the high refractoriness and adverse effects of HCHs on the ecosystem and human beings [3, 15, 16], several HCH isomers are considered persistent organic pollutants (POPs) by the Stockholm Convention [10] and classified as neurotoxic, carcinogen, and teratogen by the Environmental Protection Agency (EPA) and the World Health Organization (WHO) [17, 18]. The structure and main chemical properties of HCH isomers are given in Table 1.
\nName, CAS, water-solubility, molecular weight (MW), and chemical structure of the main HCH isomers.
HCH isomers included in the Stockholm Convention.
\n
Due to its toxicity, the production and use of lindane have been banned in most countries, including Europe and the United States [24, 25], but many landfills and the surroundings of the lindane production sites remain polluted nowadays, with soil and groundwater contaminated by these compounds [5, 6, 7, 8, 9, 10, 11, 12, 13, 14]. The low tolerance limits allowed for HCHs in water and soils have prompted a growing interest of the scientific community to develop simple, cost-effective, and fast methods for the degradation of these pollutants. Conventional methods commonly used include the excavation of polluted soil and its further containment in secure landfills. The traditional groundwater treatment consists of pump-and-treat, with adsorption in activated carbon as a common treatment. However, these options are very expensive and are not a definitive solution since the destruction of the pollutants is not achieved. Therefore, they are considered neither sustainable nor definitive remediation methods [1].
\nSome studies have focused on the remediation of HCHs in the aqueous phase, dealing with groundwater treatments applied in situ [26] or on-site [16, 27, 28]. However, only a few works are found in the literature concerning the remediation of soils contaminated by HCH wastes. The objective of these treatments is the chemical or biological degradation of HCHs. They were carried out to the remediation of soils artificially spiked with HCH isomers and soils with real HCH contamination. The chemical technologies used for the remediation of soils polluted with high HCH concentration are analyzed and discussed in the following sections. Biological treatment of these highly contaminated soils requires long times [29] [30] and are not treated here. Moreover, to better understand the particularities and characteristics of the problem of HCH polluted soils, a description of the main sites contaminated by HCH wastes is also included.
\nSoils contaminated by HCH isomers show a wide range of concentration values. In general, these soils come from areas where residues of the lindane manufacture or other chlorinated pesticides have been dumped and uncontrollably accumulated. The contamination is present in the form of particulate matter (white particles of HCH wastes in soil distinguishable with the naked eye) and/or adsorbed into the soil. When a particulate matter of HCH wastes is not reported, the concentration of 𝛃-HCH in the soil is usually higher than that of 𝛂-HCH [31], indicating that HCH isomers are adsorbed into the soil. Real soils polluted with adsorbed HCH isomers have been reported in the following works:
Riparian area of the Mulde river (Germany). In this case, two highly contaminated sites were characterized by Keller (Kel) and Spittel (Spi) [32]. The concentration of 𝛃-HCH with the depth of the soil is shown in Figure 3. It can be seen that the HCH concentration decreases with this variable.
Teltow Canal (Berlin) [33], where the concentrations of HCH isomers in the soils were analyzed. In general, the isomer with the highest concentration in the different sediments studied was 𝛃-HCH, as shown in Table 2. The lack of particulate matter of HCH wastes in these sediments could explain the higher concentration of β-HCH detected in the soil.
A contaminated area in Bitterfield (Germany) was described by Wycisk et al. [13], including an old landfill used for the discharge of HCHs and other chlorinated pesticides. The concentration of β-HCH in the soil of the old landfill was higher than the concentration α-HCH, as shown in Table 3.
A gravel pit located in the northeast of France [34], contaminated by HCH wastes dumped by the PCUK company. This company stopped the manufacture of lindane in 1974. In this place, the lindane wastes were encapsulated, but β-HCH (45 mg/kg) and γ-HCH (25 mg/kg) isomers remained in the soil. The presence of α-HCH was not described in this case.
Agricultural soils in India [29], where chlorinated pesticides were probably stored in the past. The concentration of HCH isomers in these agricultural soils varied from 4.9 to 7.4 mg/kg soil, with a concentration of α-HCH lower than that one of β-HCH.
A contaminated industrial site in Beijing, China. In this case, Liang et al. [35] collected and analyzed soil samples from an old organochlorine pesticide plant located in Beijing. The soil, which was sieved (2 mm) and homogenized, was characterized, obtaining the following parameters: pH 7.8, total organic carbon 6.2 g/kg, total nitrogen 0.82 g/kg, and moisture 2.8%. The soil contained α-HCH, β-HCH, p,p′-DDT, or p′-DDT, p,p′-DDE, and p,p′-DDD with concentrations of 2.7, 10.8, 12.9, 3.1, 2.0, and 2.6 mg/kg, respectively.
Farm soils in Kazakhstan, where obsolete pesticides were stored during decades [36]. As can be seen in Table 4, the concentration of β-HCH was higher than the concentration of α-HCH.
Soil contamination by β-HCH as a function of depth [32].
Depth, cm | \nα-HCH | \nβ-HCH | \nγ-HCH | \nδ-HCH | \nε-HCH | \n∑HCH | \n
---|---|---|---|---|---|---|
00–10 | \n17 | \n120 | \n41 | \n64 | \nb.d.l. | \n242 | \n
10–15 | \n48 | \n110 | \n29 | \n44 | \nb.d.l. | \n231 | \n
15–20 | \n61 | \n140 | \n42 | \n68 | \nb.d.l. | \n311 | \n
25–30 | \n79 | \n170 | \n50 | \n97 | \nb.d.l. | \n396 | \n
55–60 | \n120 | \n65 | \n47 | \n130 | \nb.d.l. | \n362 | \n
60–65 | \n28 | \n20 | \n12 | \n40 | \nb.d.l. | \n100 | \n
65–70 | \n18 | \n13 | \n7.6 | \n25 | \nb.d.l. | \n64 | \n
95–100 | \n0.3 | \n0.3 | \n0.2 | \n2.1 | \nb.d.l. | \n3 | \n
Concentration (μg/kg) of HCH isomers in Teltow Canal sediments, Berlin [33].
b.d.l = below detection limit
Depth, cm | \nα-HCH | \nβ-HCH | \nγ-HCH | \nδ-HCH | \n
---|---|---|---|---|
00–10 | \n440.0 | \n702.5 | \n23.3 | \n10.3 | \n
10–20 | \n535.6 | \n574.3 | \n9.5 | \nb.d.l. | \n
20–30 | \n109.0 | \n60.1 | \n0.4 | \nb.d.l. | \n
30–60 | \n0.5 | \n6.4 | \n0.3 | \nb.d.l. | \n
60–80 | \n0.4 | \n2.4 | \n0.1 | \nb.d.l. | \n
80–100 | \n0.2 | \n7.0 | \nb.d.l. | \nb.d.l. | \n
HCH content (mg/kg) in vertical floor profiles on Spitelwasser [13].
b.d.l = below detection limit
Hot points | \n𝛂-HCH | \n𝛃-HCH | \n𝛄-HCH | \n
---|---|---|---|
MAC | \n0 | \n100 | \n100 | \n
Point 1 | \n67.1 ± 9.1 | \n176.0 ± 23.3 | \n22.2 ± 3.2 | \n
Point 2 | \n15.3 ± 7.3 | \n83.2 ± 5.5 | \n13.0 ± 4.2 | \n
Point 3 | \nb.d.l. | \nb.d.l. | \nb.d.l. | \n
Concentration values of HCH isomers (μg/kg) in various soils analyzed [36].
b.d.l = below detection limit
On the other hand, other works dealing with soils contaminated by HCH wastes reported the presence of white granules of particulate HCHs noticed with the naked eye. In these studies, the average concentration of α-HCH measured in the soil was higher than that of β-HCH, which agrees with the composition of technical-HCH. These studies are listed below:
Contaminated soils in Galicia (Spain) affected by the industrial activity of lindane production [37]. The soils were analyzed at different points and depths, and the concentration values of HCH isomers found ranged from 5 to around 80,000 mg/kg, with very different concentrations depending on the depth analyzed. The highest values of HCHs (81,035 mg/kg) probably corresponds to the presence of HCH isomers in the form of particulate matter. Fragments and dust of this white substance were present in the studied area. This material was also analyzed to determine the presence of technical-HCH wastes. The technical HCH produced in that fabric presented the following composition: 77% of α-HCH, 16% of β-HCH, 5% of γ-HCH, and 2% of δ-HCH. The composition of some soil samples analyzed in this work is shown in Figure 4. It was noticed that samples with the highest proportion of β-HCH correspond to those with the lowest total concentration of HCHs, whereas samples with the highest percentage of α-HCH correspond to soils with higher total HCH concentration (probably as grains of technical-HCH wastes). The presence of HCH in the form of particulate matter could add difficulties in the remediation of these sites.
Sabiñánigo (Huesca, Spain), with two landfills contaminated with HCH wastes dumped by INQUINOSA, a lindane factory which operated from 1975 to 1988 [11]. High concentrations of HCHs in the soil were measured (Table 5), and a higher concentration of α-HCH isomer than that of β-HCH was reported, which is in agreement with the presence of HCHs as particulate matter detected in that soil.
City of Meninos, Brazil, where contaminated soils were found near to a former lindane factory, which operated between 1950 and 1962 [12]. Although the distribution of HCH isomers in the soil was not reported, the high concentration of these pollutants measured (several thousand mg/kg), indicates the presence of HCH-wastes as particulate matter.
Santo André’, Sao Paulo (Brazil) with HCH-contaminated land [38]. In this study, there is no explicit indication of the presence of HCH particulate matter, but the high proportion of α-HCH, shown in Figure 5, seems to confirm this hypothesis.
Soil contaminated by HCH wastes in the Midwest (USA). Phillips et al. [30] studied three areas (A, B, and C) with a high concentration of HCHs, mainly due to the presence of HCHs granules. The total concentration of HCHs along a west to east gradient ranged from 22,430 to 1069 mg/kg in the A zone and from 21,100 to 730 mg/kg in the B zone, whereas in zone C, the concentration ranged from 52 to 1427 mg/kg. The composition of HCH wastes was rich in the isomer α (α-HCH 20,000 mg/kg and β-HCH 2000 mg/kg).
Percentages of HCH isomers found in soil samples [37].
\n | α-HCH | \nβ-HCH | \nγ-HCH | \nδ-HCH | \nε-HCH | \n∑HCH | \n
---|---|---|---|---|---|---|
Maximum | \n57,000 | \n5600 | \n9700 | \n2200 | \n2700 | \n74,730 | \n
Mean | \n2303.2 | \n245.5 | \n406.8 | \n105.7 | \n138.9 | \n3200.2 | \n
HCH concentration (mg/kg) in soils located at Sabiñánigo landfills [11].
Distribution of HCH isomers in soils located at Santo André’ and Sao Paulo (Brazil) [38] and in technical HCH.
Physicochemical treatments have been studied in the literature applied to real or spiked contaminated soils. The last ones obtained by contacting the soil with concentrated solutions of HCHs solved in different organic solvents and the subsequent evaporation of the solvents. In the first case (real soils), the contaminants can be absorbed into the soil (a higher concentration of the isomer β-HCH is noticed) or present as HCH granules (with a higher concentration of the isomer α-HCH). The proportion of HCH isomers found also depends on the composition of the dumped HCH wastes or the spiking procedure (in the case of spiked soils).
\nIn general, β-HCH is always the most recalcitrant isomer regardless of the treatment tested (biological or chemical oxidation and biological or chemical reduction). The following sections summarize the works found in the bibliography related to physicochemical remediation treatments.
\nThermal treatments have been traditionally applied to the remediation of soils contaminated with persistent organic pollutants, as HCH wastes. However, the use of high temperatures has major drawbacks, such as the low-cost effectivity of the process and the generation of compounds even more toxic than the starting ones when chlorine is in the structure of the organic pollutant, such as dioxins and furans. The main thermal treatments found in the literature for the remediation of HCH-polluted soils are described below:
\nThis thermal process was applied to the remediation of real soils located in Sao Paulo (the distribution of HCH isomers suggests the presence of HCHs in the form of particulate matter) [38]. The excavated soils were subjected to high temperatures (up to 450°C), as is indicated in the scheme of the heat treatment plant shown in Figure 6. The results obtained for the abatement of the different HCH isomers are shown in Figure 7, as a function of the reaction time and the temperature of the treatment. An important degradation of HCHs is achieved in only a few hours of reaction, although temperatures above 250°C are required. It should be noted that β-HCH is also the most recalcitrant isomer.
\nHeat treatment plant for HCH-contaminated soils [38].
Removal of HCHs from the soil under different heat treatment conditions [38].
Rozdyalovskaya and Chekryshkin [39] studied the destruction of pure lindane at high temperature by using basic catalysts. The deep oxidation of lindane on a catalyst can be represented by the following reaction (Eq. (1)):
\nThe oxidation was performed in the temperature range of 400–750°C in the presence of some fused catalysts. The catalyst samples were prepared by dehydration, weighing, and mixing of the components and liquid phase synthesis (melting).
\nThe highest activity in the reaction of deep oxidation of lindane was obtained using a molten catalyst based on a eutectic mixture of carbonates of alkali metals with 10 wt % of V2O5 and CuO. Moreover, simultaneously with the reaction of deep oxidation of lindane, its dehydrochlorination in a melt of sodium and potassium hydroxides was also noticed. The temperature required for lindane destruction in this process was higher than 450°C. Although there is no mention about dioxins and furans in the work, when working at these temperatures, these compounds are usually generated.
\nThis treatment consists of the addition of specific reagents to the soil contaminated with halogenated organic compounds under strong temperature conditions. The process of dehalogenation is achieved by replacing halogen atoms or by the decomposition and partial volatilization of the contaminants [40]. Among dehalogenation processes, base-catalyzed decomposition (BCD) and alkali glycol/polyethylene glycol (APEG) processes [41, 42, 43] can be considered. These treatments have been successfully applied to remediate soils and sediments contaminated with chlorinated organic compounds, especially PCBs, dioxins, and furans, but high temperatures are required (150–330°C).
\nDifferent oxidants and activators have been tested in the treatment of soils with real or simulated HCH contamination. However, no studies have been found to date on soils contaminated with HCHs in the form of particulate matter. The presence of this kind of pollution (particulate matter) could pose an additional limitation since the prior solubilization of these granules would be necessary. This phenomenon would be controlled by the interfacial surface between water and the solid phase. The contact between the two phases will increase (i) as the particle size of the HCH granules decreases and (ii) the agitation of the slurry soil-aqueous phase increases.
\nThe main results obtained in the remediation of HCH-contaminated soils with oxidation technologies are described below:
\nPeng et al. [44] tested the thermal activation of persulfate (PS), at 20 and 40°C, in the treatment of soils artificially contaminated with 800 mg/kg of lindane (γ-HCH was the only HCH isomer studied in this work). The water/soil mass ratio selected was 4:1, and the concentration of PS in the aqueous phase was 0, 5 and 50 g/L. At 20°C, there was no reaction noticed, whereas at 40°C, lindane was eliminated with 50 g/L of PS in 15 days reaction time. When a lower concentration of PS was used (5 g/L) at the same temperature (40°C), the reaction extent was small, as shown in Figure 8.
\nRemoval of lindane from soil after 10 and 20 days with temperature-activated PS at 40°C and using different concentrations of PS [44].
These authors also used the alkaline activation of PS. When this treatment was applied, it was observed that lindane was converted into trichlorobenzenes [44]. Regrettably, there is no information about how these compounds (trichlorobenzenes, TCBs), or other reaction by-products, disappear once formed. The authors did not study the abatement of other HCH isomers than γ-HCH. Recently, Dominguez et al. [45] studied the oxidation of real soils polluted with α-HCH (120 mg/kg) and β-HCH (35 mg/kg) isomers by persulfate activated by alkali finding that the hydrolysis of the β-HCH was the limiting step and that the oxidation rate of TCBs increases notably when the reaction temperature rises from 20 to 40°C.
\nUsman et al. [34] used artificially contaminated soils (100 mg/kg of each HCH isomer: α, β, γ, δ in sand) and real contaminated soils (concentration of β-HCH = 45 mg/kg and γ-HCH = 25 mg/kg). The high concentration values of β-HCH in the real soil indicate that no particulate matter was present in that soil. The oxidation treatments tested by the authors were persulfate activated by temperature, Fenton reagent (H2O2 + Fe), and permanganate. The carbonate content in the real soil was relatively high (195 g CO3Ca/kg soil), which is relevant for the potential application of H2O2 as an oxidant, since it would lead to high unproductive consumption of the oxidant. Moreover, the pH of the soil was slightly alkaline (8.05), which also hinders the application of iron as an activator due to its precipitation at this pH.
\nA water/soil mass ratio = 20:1 and a large excess of oxidant (17 g/L H2O2 and 71 g/L PS) were used, with molar ratios of Fe/H2O2 = 1/10 and Fe/PS = 1:2. The higher proportion of Fe used in the activation of PS than in the Fenton process is due to the fact that in the first case, iron is a reagent that is consumed with the progress of the reaction (Eq. (2)), whereas in the case of Fenton reagent, iron is a catalyst, which is continuously regenerated during the radical species production. Fe(II) reacts with hydrogen peroxide to give hydroxyl radicals and Fe(III) (Eq. (3)), which is after regenerated to Fe(II) reacting with another molecule of hydrogen peroxide (Eq. (4)). The reaction of HCHs with the radical species generated by both processes,\n
In the aforementioned work, the authors compare the results obtained with the following treatments: H2O2 only, Fenton reagent (H2O2 + Fe (II)), PS only, PS activated with Fe(II) and potassium permanganate after 24 hours [34]. When iron is used, it is necessary to carry out the reaction at acid pH to avoid iron precipitation, which results unaffordable in the case of soils with high carbonate content. The results and specific conditions obtained in each treatment for both spiked (a) and real contaminated (b) soils are shown in Figure 9. The most recalcitrant HCH isomer was β-HCH regardless of the treatment tested, and the best results were obtained with Fe-activated PS (it should be noted that it was necessary to bring the pH to the acidic zone 2–3). Data about the consumption of the different oxidants are not supplied in the article, but it is expected that H2O2 reacted unproductively when this oxidant is applied to the remediation of the real polluted soils (Eq. (7)), being the reason for the cause of the lower HCH conversion obtained with this treatment.
\nDegradation of HCH isomers in (a) spiked sand with HCHs and (b) real contaminated soil [34].
García-Cervilla et al. have recently studied the remediation of a soil located at 14 m below the ground level in an alluvial of an old landfill contaminated with liquid wastes of lindane production [46] at Sabiñanigo (Spain). A high carbonate concentration was also found in this soil (>45%), and the alkaline activation of PS was selected as a remediation technology. The organic and inorganic composition of the soil sieved at two particle sizes: F (dp < 0.25 mm) and G (0.25–2 mm) is summarized in Table 6. This soil presented high HCH concentration in some points (up to 9000 mg/kg) due to the adsorption of DNAPL (dense non aqueous phase liquid) that percolated through the soil and reached the alluvial. The absence of the isomer β-HCH in the DNAPL is the reason of the lack of this HCH isomer in the soil studied.
\n\n | F, dp < 0.25 mm | \nG 0.25 < dp < 2 mm | \n
---|---|---|
TOC, mg/kg | \n2820 | \n840 | \n
TC, mg/kg | \n54,660 | \n54,840 | \n
Carbonates (as CaCO3) (%w) | \n43.2 | \n45.0 | \n
Fe, mg/kg | \n33,078 | \n31,662 | \n
∑HCH, mg/kg | \n6597.3 | \n1735.5 | \n
∑Heptachlorocyclohexanes, mg/kg | \n1997.0 | \n690.7 | \n
Total mg/kg | \n10,109 | \n3346 | \n
Inorganic and organic composition of polluted soil (14 mg g l) [46].
As previously commented, persulfate activated by alkali was applied for the remediation of this soil. This method follows a free radical mechanism [47, 48, 49], summarized in Eqs. (8) and (9).
\nThe addition of an alkali provoked that HCH and heptachlorocyclohexane isomers adsorbed into the soil as a residual phase were converted to trichlorobenzenes and tetrachlorobenzenes, respectively, in less than 48 h. The dehydrochlorination reactions at alkaline conditions, shown in Figure 10, were previously described elsewhere [50, 51, 52].
\nDehydrochlorination reactions at alkaline pH [52].
At pH above 12, it has been noted that hydroxyl radicals (OH∙, E0 = 2.7 V) are predominant against sulfate radicals (SO4∙−, E0 = 2.6 V) [53]. In addition to hydroxyl radical, superoxide radical is also produced in the alkaline activation of persulfate, as can be seen in Eq. (8). These species are capable of producing a nucleophilic substitution when reacting with halides, as described in Figures 11 and 12, where trichlorobenzene is mineralized by the attack of both superoxide [54] and hydroxyl [44] radicals.
\nNucleophilic substitution of the superoxide radical in the reaction with trichlorobenzene as an example of oxidation reaction. Tetrachlorobenzene isomers follow the same reaction mechanism [54].
Attack of hydroxyl radicals on trichlorobenzene as an example of oxidation reaction. Tetrachlorobenzene isomers follow the same reaction mechanism [44].
The main results obtained in this work are shown in Figure 13 [46]. The molar ratio NaOH/PS ratio was 2:1, the mass ratio water/soil was 10:1, and the concentration of PS varied between 25 and 100 g/L. As can be seen, more than 1 month was required for the remediation of the soil with a particle size <0.25 mm, due to the high pollutant concentration and the strong adsorption of the pollutants to the soil (higher presence of clays than the other fraction) found in the fraction “F” .
\nConversion of isomers: (a) 1,2,3 TCB, (b) 1,2,4 TCB, (c) TetraCBs-a in soil F, and (d) TetraCBs-b in soil F and G after 509 h. \n\n\n\n\nC\n\n\nP\nS\n\n\n/\n\n\nC\n\n\nN\na\nO\nH\n\n\n=\n1\n\n\n. Soil F: Diameter lower than 0.25 mm, soil G: Diameter between 0.25 and 2 mm [46].
Muñoz Morales et al. [55] studied the remediation of a soil artificially contaminated with lindane (100 mg/kg). An anionic surfactant, SDS, was used to extract the pollutant from the soil (0.1 g SDS per g soil) using a liquid to solid phase mass ratio of 10. Subsequently, lindane extracted from the soil and solved in the aqueous phase was removed by electrooxidation. Therefore, this was a treatment train consisting, firstly, in the solubilization of the pollutant and secondly, in the selective oxidation of the pollutant, in aqueous emulsion. A diamond electrode was used to in-situ generate hydrogen peroxide by injecting air. Remediation times of 400 min were needed, and the surfactant was recycled for further washing cycles.
\nZero-valent iron (ZVI) has attracted the interest of the scientific community over the past decade for its potential to remediate a wide variety of environmental contaminants both in superficial and groundwater [56]. The use of ZVI over other metals is a preferred choice due to its high abundance, low cost, and benign environmental impact [25, 56, 57]. Among other pollutants, ZVI showed high efficiency in the treatment of chlorinated organic compounds such as HCH isomers [2, 4, 17, 25, 26, 57, 58, 59, 60, 61, 62, 63, 64, 65].
\nMost of the reported works are focused on the degradation of lindane and the use of ZVI nanoparticles [2, 3, 4, 17, 25, 58, 60, 61, 62, 63] or the combination of ZVI with other metals, Pd being the most studied [17, 25, 60]. In the presence of ZVI nanoparticles, lindane can be eliminated in 24 hours reaction time when this pollutant is dissolved in water [2] or present in spiked soils [4]. In the case of using bimetallic Pd-Fe nanoparticles [17, 25, 60] or more complex systems, like carbon-supported Cu-ZVI nanoparticles [3] or carboxymethylcellulose Fe/Ni nanoparticles [63], the reaction times for lindane dichlorination can be even decreased. It has been reported that anaerobic conditions favor lindane degradation in the presence of stabilized iron nanoparticles [17], and the temperature has a beneficial effect on the pollutant degradation rate [60], whereas lindane degradation decreases with pH increasing, initial lindane concentration, and in the presence of cations [60]. Several degradation pathways for lindane degradation have been proposed based on the detection of certain reaction intermediates during lindane dichlorination reactions in the presence of ZVI nanoparticles [3, 25, 62, 63].
\nAlthough encouraging results in HCH treatment in the presence of these materials have been achieved, the low stability of iron nanoparticles due to aggregation [3, 17, 25, 62] and the unaffordable cost of noble metals like Pd [57] has encouraged the use of ZVI in the form of microparticles during the last years, with lower cost and higher stability [26, 64, 65].
\nThe predominant mechanism for the degradation of lindane using ZVI is the reductive dehalogenation of the pollutant, owing to the electron exchange between the HCH molecule and zero-valent iron [3, 58, 60, 62, 65]. Benzene is obtained as the final product of lindane reduction (along with chlorides), as is shown in Figure 14.
\nDechlorination pathway of lindane over zero-valent iron microparticles [65].
As occurred with chemical oxidation, 𝛃-HCH presents high recalcitrance towards chemical reduction, in both aqueous and soil phases [2, 65, 66] due to the chlorine’s position and the low water solubility of this HCH isomer.
\nEven though promising results have been obtained with ZVI in the degradation of HCHs in the aqueous phase, the use of this material for soil remediation entails additional problems. The application of ZVI in the form of microparticles on contaminated soils would yield low HCH conversion due to the hindered contact between the solid phases (soil and ZVI microparticles). Using ZVI nanoparticles for soil remediation is limited by problems of agglomeration and the high cost associated. Furthermore, if HCH granules are present in the soil, a remarkable decrease in the efficiency of the dechlorination treatment is expected, due to the expected poor contact between the two solid phases.
\nIn this technology, surfactants are used to solubilize the contaminants absorbed into the soil in the aqueous phase. The resulting solution requires a second stage in which the objective is the selective oxidation of the contaminant from the emulsion and the surfactant recover for a next use [67, 68]. However, the solubilization of pollutants from solid phases is hindered by the pollutant transport from the soil to the aqueous phase.
\nThere are few papers in the literature using this technology for the treatment of HCH-contaminated soils, none dealing with the presence of particulate matter, and only spiked soils with HCHs or DNAPL were used.
\nMuñoz-Morales et al. [55] used soil washing as a first stage to remediate a soil spiked with lindane. For that purpose, the authors used an anionic surfactant, SDS. This surfactant was selected because the next step of the remediation treatment was the electrochemical oxidation of the pollutant in the emulsion, and high conductivity of the solution is required in this oxidation treatment. Using a surfactant concentration of 10 g/L in the aqueous phase, the concentration of lindane found in the aqueous emulsion was 10 mg/L, which was further oxidized by the electrochemical treatment.
\nDominguez et al. combined soil flushing (with a nonionic surfactant) and Fenton oxidation [67]. A nonionic commercial surfactant (E-Mulse 3®) was used to extract most of the residual DNAPL in the soil at column conditions. The resulting surfactant flushing solution showed a high concentration of chlorinated organic compounds (COCs = 3693 mg/L, 40% of this amount corresponded to HCH isomers, although β-HCH was not in the mixture). This emulsion was treated by the Fenton process using different concentrations of hydrogen peroxide (200%, 100%, and 50% of the theoretical stoichiometric amount for the complete mineralization of the COCs) and a molar ratio of H2O2:Fe = 32. A degradation of COCs >80% was obtained using a concentration of H2O2 ≥ 100% of the stoichiometric amount. HCHs (and other nonaromatic COCs) were less prone to oxidation by hydroxyl radicals than chlorobenzenes. The surfactant was recovered at the end of the treatment for further flushing steps.
\nRegrettably, there are no studies in the literature dealing with soil washing of soils polluted with β-HCH, the least soluble and the most stable HCH isomer against oxidation and reduction. Moreover, if the contamination of the soil by HCHs involves also the presence of particulate matter, transport resistances will be more limiting, and the step of soil washing will slow down. In this case, a good agitation or ultrasound application will be required to improve the contact between the phases and, therefore, to improve the efficiency of the process.
\nSoil contamination by the solid residues generated from the manufacture of lindane, a chlorinated organic pesticide whose use and production has been prohibited, is a great environmental problem, ubiquitous and persistent, given the high toxicity and low biodegradability of these residues in the environment. These soils contain a mixture of HCH isomers, mainly α and β, isomer β being the most recalcitrant to both chemical and biological treatments, due to its lower water solubility and higher chemical stability. This kind of contamination appears as solid HCH particles mixed with soil (usually with a higher concentration of the isomer α-HCH) or adsorbed onto the soil (with a higher concentration of the isomerβ HCH) reaching values up to several hundreds of mg HCH/kg soil. It represents a serious problem due to the large volume of wastes to be treated. Among the physicochemical treatments used, thermal processes are the traditional ones but the less sustainable because the requirement of high temperatures and, therefore, the associated costs are prohibitive for treating large amounts of wastes. In the last decade, chemical treatments have shown promising results. Amon them, oxidation with Fenton reagent or activated persulfate seems to be more suitable than reduction using zero-valent iron particles, because of the greater limitations for the contact between phases in the last one treatment. The selection of the most suitable oxidation method will depend on the type of soil (presence of carbonates and pH). On the other side, the time and method of contact will also be strongly influenced by how the contamination is present (in the form of particulate or adsorbed matter).
\nThe authors acknowledge the financial support from SARGA, a public company of the Government of Aragon (project ref. 5507001-182 funded), from the Regional Government of Madrid, through the CARESOIL project (S2018/EMT-4317), and by the Spanish Ministry of Science (project CTM2016-77151-C2-1-R). The authors thank SARGA and the Department of Climate Change and Environmental Education, Government of Aragon, for their support during this work.
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\n\nThe Social Media Community Manager and Marketing Assistant will report to the Senior Marketing Manager. They will work alongside the Marketing and Corporate Communications team, supporting the preparation of all marketing programs, assisting in the development of scientific marketing and communication deliverables, and creating content for social media outlets, as well as managing international social communities.
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