三峡库区消落带汞的形态转化与释放特征
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摘要
汞及其化合物(特别是烷基汞)具有很强的生物毒性、较快的生物富集放大速率和较长的脑器官生物半衰期。自上世纪50年代以来,环境汞污染问题一直受到人们的广泛关注。而去年召开的第10次“汞作为全球污染物(Mercury As a Global Pollutant)”的国际会议议题(2011年7月,加拿大哈利法克斯)以及自2010年开始的全球汞控制公约政府间多轮谈判表明,作为“全球性污染物”的汞及其产生的环境问题,仍是目前乃至今后较长一段时间内人类必须面对的热点问题。
水库是一个典型的“汞敏感生态系统”,水库环境汞环境化学行为的研究至关重要。三峡水库属特大型年调节水库,按照三峡水库“蓄清拍浊”的调度方式,水库正常蓄水位为175m,防洪限制水位为145m。这种水库调度方式使得水库周边形成垂直高度为30m、面积约350km2的水库消落带。一方面,由于是库区径流的汇集地带,消落带自然也成为环境汞的汇集区(汞汇)。另一方面,它所汇集的汞等污染物最终又会影响到水体质量(汞源)。同时,周期性的水位变化使得消落带将长期处于干湿交替的变换过程,消落带沉积物和上覆水体的性质将会发生一系列变化,如pH、Eh、电导率、微生物、有机质等,沉积物和水体环境的变化对汞的迁移转化以及水生生物的富集将有很大影响。
目前,对三峡库区汞问题的研究还仅限于对水体汞含量、鱼汞分布、淹没土壤汞含量的调查,而对消落带沉积物中汞的赋存形态转化规律与机制、界面释放特征及其生态效应等方面的研究还相当缺乏。基于此,本文采用野外现场调查采样分析与室内模拟实验研究相结合的方法,对三峡库区消落带沉积物汞在沉积物-大气界面、水体-大气界面、土壤(沉积物)水界面的释放特征,及沉积物中汞赋存形态、转化规律与机制进行了系统研究。
结果表明,在水位涨落的不同时期,库区消落带土壤(沉积物)-大气界面汞交换通量也是不同的。11月份土壤(沉积物)-大气界面汞交换通量最高,为28.17±36.17ng/m2h。1月份库区消落带土壤(沉积物)-大气界面汞交换通量最低,为-6.80±12.35ng/m2h,总体表现为沉降。总体来看,三峡库区消落带土壤(沉积物)-大气界面汞的交换通量高于国内外森林土壤和背景土壤汞的交换通量,但低于国内外对湿地沉积物、冲积平原土壤及农田土壤的释汞通量。三峡库区消落带土壤(沉积物)-大气界面汞的交换通量具有一定的日变化特征,晴天白天从早晨开始士壤(沉积物)-大气界面汞的交换通量逐渐增加,到正午左右达到峰值,之后交换通量逐渐减小。夜间土壤(沉积物)-大气界面汞交换通量的变化较为平稳,大部分数值在0ng/m2h附近波动。而阴天库区消落带土壤(沉积物)-大气界面汞交换通量变化的趋势不明显。暖季白天土壤(沉积物)-大气界面汞交换通量都以释放(正值)为主,冷季白天则以沉降(负值)为主。夜间,除11月份表现为释放外,其余三个采样时段夜间土壤(沉积物)-大气汞交换总体上均表现为沉降。光照强度是影响白天土壤(沉积物)-大气界面汞交换的最主要因素,随着光照强度的增强,土壤(沉积物)释汞通量呈线性(11月)、对数(1月)或乘幂(5月、7月)形式增加。温度也是影响库区消落带土壤(沉积物)-大气界面汞交换的重要因素,土壤(沉积物)温度和气温都与土壤(沉积物)-大气界面汞释放通量呈指数关系。大气汞浓度是影响消落带夜间土壤(沉积物)-大气界面汞交换的最主要因素。根据实测数据得到了基于诸影响因素的库区消落带土壤(沉积物)-大气汞交换通量的逐步回归方程,并据此估算了三峡库区消落带土壤(沉积物)-大气界面平均汞交换通量为12.77ng/m2h,库区消落带沉积物向大气释汞量为19.53kg/a。
库区消落带水位涨落的不同时期,水体-大气界面汞交换通量也是不同的。7月份水体-大气界面汞交换通量最高,为24.47±36.32ng/m2h,是其他采样时段水体-大气界面汞交换通量的约4-6倍。11月份库区消落带水体-大气界面汞交换通量最低,为3.83+18.66ng/m2h,最大沉降值(负值)还是出现在冷季的1月份。总体来看,三峡库区消落带水体-大气界面汞的释放通量与国内外学者对海洋、河口水域、水库、湖泊水体的释汞通量相近,但低于国外对大型湿地水体释汞通量的监测结果。三峡库区消落带水体-大气界面汞交换通量的具有一定的日变化特征,从早晨开始水体-大气界面汞的交换通量逐渐增加,到正午左右达到峰值,之后交换通量逐渐减小。夜间水体-大气界面汞交换通量的变化较为平稳,大部分数值在Ong/m2h附近波动。无论是暖季还是冷季,白天水体-大气界面汞交换通量都以释放(正值)为主。夜间,冷季库区消落带水体-大气界面汞交换通量以沉降(负值)为主,暖季界面汞的释放和沉降达到平衡。光照强度是影响白天水体-大气界面汞交换的最主要因素,随着光照强度的增强,水体释汞通量呈对数(11月、1月、5月)或乘幂(7月)形式增加。温度也是影响库区消落带水体-大气界面汞交换的重要因素,水温和气温都与水体-大气界面汞释放通量呈指数关系。大气汞浓度是影响消落带夜间水体-大气界面汞交换的最主要因素。根据实测数据得到了基于诸影响因素的库区消落带水体-大气汞交换通量的逐步回归方程,并据此估算了三峡库区消落带水体／大气平均汞交换通量为14.45ng/m2h,库区消落带水域向大气释汞量为22.21kg/a。
库区消落带沉积物总汞(THg)含量略高于河流沉积物汞的背景值,但比其他被汞污染的沉积物低很多,三峡库区消落带汞的污染程度较轻。7月份和9月份采集的库区消落带沉积物甲基汞(MeHg)的含量分别为0.128±0.028ng/g和0.031±0.027ng/g。7月份MeHg:THg的值为0.296±0.154%,9月份MeHg:THg的值为0.069±0.081%,两者相差近3倍。不同类型沉积物MeHg：THg的值也存在差异,总体表现为淹水沉积物>半淹水沉积物>落干沉积物。但总体来讲,三峡库区消落带沉积物中汞甲基化的速率和程度还处于较低的水平。7月份库区消落带沉积物中汞的赋存形态依次为强络合态(46.99%)、硫化物结合态(36.28%)、有机及其他络合态(9.57%)、水溶态(4.49%)、胃酸溶态(2.67%)。强络合态是7月份库区消落带沉积物中汞主要的赋存形态,即7月份库区消落带沉积物中汞主要是与铁锰氧化物、无定形有机硫化物等的络合物,以及元素汞和结合于矿物质晶格结构中的汞。9月份库区消落带沉积物中汞的赋存形态依次为硫化物结合态(46.47%)、强络合态(34.73%)、有机及其他络合态(14.18%)、胃酸溶态(2.40%)、水溶态(2.21%)。硫化物结合态是9月份库区消落带沉积物中汞主要的赋存形态,即9月份库区消落带沉积物中汞主要是与硫化物结合态的汞。伴随着淹水沉积物的出露,由于无机汞在相态分布中的溢出效应,强络合态的所占的比例逐渐下降,水溶态和胃酸溶态所占的比例逐渐升高。沉积物落干过程中,经历波浪和坡面径流对库岸沉积物的冲刷,导致弱结合态汞的流失,而强结合态汞如硫化物结合态的汞的比例的增加。室内模拟实验的结果也证实了沉积物在出露过程中,确实存在无机汞相态分布的“溢出效应”。即伴随着沉积物含水量的下降,强络合态的所占的比例逐渐下降,水溶态和胃酸溶态所占的比例逐渐升高。淹水沉积物中汞的迁移性要高于落干沉积物,沉积物出露过程中汞的迁移性增加,落干沉积物淹水后汞的迁移性也会增加。
淹水期间浅水沉积物(处理A)中甲基汞占总汞的比例为0.41+0.29%,深水沉积物(处理B)中甲基汞占总汞的比例为0.74±0.52%,两处理沉积物甲基汞占总汞的比例之间存在显著性差异(p=0.009),也就是说深水沉积物比浅水沉积物具有更高的甲基汞产率。处理A与处理B上覆水体中活性汞、溶解态汞、甲基汞含量之间没有显著性差异,沉积物孔隙水中活性汞、溶解态汞、甲基汞含量之间也没有显著性差异。但上覆水和沉积物孔隙水活性汞、溶解态汞、甲基汞含量存在显著的差异,也就是说活性汞,溶解态汞,甲基汞在沉积物孔隙水与上覆水之间存在浓度梯度,沉积物孔隙水中活性汞,溶解态汞,甲基汞有向上覆水中迁移扩散的趋势。不同淹水时间两处理沉积物孔隙水中活性汞的扩散通量分别为12.79±5.06ng/m2d(处理A)和13.24+3.68ng/m2d(处理B)；溶解态汞的扩散通量分别为154.65±47.12ng/m2d(处理A)和160.23±56.19ng/m2d(处理B)。甲基汞的扩散通量分别为7.61±3.39ng/m2d(处理A)和7.79±4.56ng/m2d(处理B)。各处理之间沉积物／水界面间活性汞,溶解态汞,甲基汞扩散通量没有显著性差异。库区消落带孔隙水中的活性汞对上覆水体活性汞含量的贡献率为0.002-0.011%,孔隙水中的溶解态汞对上覆水体溶解态汞含量的贡献率为0.002-0.017%,孔隙水中的甲基汞对上覆水体甲基汞含量的贡献率为0.015-0.143%。
Mercury and its compounds (especially alkyl mercury) have very strong biological toxicity. And the rate of biological enrichment and amplification were higher than other heavy metals. Meanwhile, their biological half-lifes in brain organ were also much longer. The problem caused by mercury pollution has been aroused public concern. The successful holding of10th International Conference on Mercury as a Global Pollutant (ICMGP) in Halifax, Nova Scotia, Canada in July24-29,2011. And the Intergovernmental Negotiation Committee (INC) organized by UNEP Ad Hoc Open Ended Working Group on Mercury has also been held three times since2010. All of these events do suggest that the focus of international research on mercury have been rise gradually in recent years.
The ecosystem of reservoir was very sensitive to mercury. It is an indisputable fact that the mercury concentration in fish will higher than other water body after the reservoirs were built. It has already become one of hot topics of mercury research. The largest reservoir in china, the famous Three Gorges Reservoir, is located at the middle reach of the Yangtze River. The water level of the Three Gorges Reservoir fluctuates from145m in summer (May-September) to175m in winter (October-April), and the water-level-fluctuation zone (WLFZ) with area of350km2were formed. On the one hand, WLFZ is an effective sink for mercury due to its integration and accumulation of mercury from both the terrestrial and aquatic environments (the surface water of adjacent river). And mercury from industrial activities and domestic consumption in the uplands may be accumulated in the WLFZ by flooding. On the other hand, the higher Hg concentration in sediment will also influence the quality of overlying water.
Periodic fluctuation of reservoir water caused the sediment could face a prolonged period of dry and wet alternate. And there will be a series of physical, chemical, mechanical changes (such as pH, Eh, electrical conductivity, microorganisms, organic matter etc.) in sediment and overlying water. There is, however, something that eventually will have a much bigger impact on Release, Movement and Transformation of Mercury in Water-Level-Fluctuating Zone of Three Gorges Reservoir Area, China.
Up to now, the research about mercury problem in WLFZ of Three Gorges Reservoir Area were just confined to Hg concentrations in water, the distribution of Hg in fish body and investigate the Hg concentrations in flooding soil. However, the speciation of mercury and its transformation in sediment in WLFZ, the characteristic of release from interface and its ecological effects were very insufficient. It's exactly because of this, our research were carried out based on field investigation, field sampling, and laboratory experiments. In this thesis, the author attempts to make a systematic analysis of mercury exchange fluxes on water/air, sediment/air and sediment/water interface in WLFZ. And the speciation of mercury and its transformation in sediment in WLFZ were also included.
The result showed that the mercury exchange fluxes on soil (sediment)/air surface were difference with water level fluctuated in different sampling periods. The highest value of mercury exchange fluxes on soil (sediment)/air surface was28.17±36.17ng/m2h (emission, November), and the minimum value of mercury exchange fluxes on sediment/air surface was-6.80±12.35ng/m2h (deposition, January). In general, the mercury exchange fluxes on soil (sediment)/air surface in WLFZ were higher than emission from forest soil and background soil at home and abroad, and lower than emission from sediment in wetland, soil in floodplain and soil in agriculture field. There were trend of diurnal variation characteristics of mercury exchange on soil (sediment)/air surface in WLFZ in sunny day. At the daytime, a sharp increase in flux starting in the morning and peak emission at midday, then decreased gradually with sunlight decreased. At the nighttime, mercury exchange fluxes on soil (sediment)/air surface in WLFZ have been stabilized and its value remained at0ng/m h fluctuation. There were non-significant trend of diurnal variation characteristics of mercury exchange on soil (sediment)/air surface in WLFZ in cloudy day. Mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by emission at daytime in warm season. On the contrary, mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by deposition at daytime in cold season. Except November, mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by deposition at nighttime.
Solar radiation was the most important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ at daytime. The regression analysis reveals that mercury emission from soil (sediment)/air surface increased line (November), logarithmic (January) and power (May and July) with raised solar radiation. Temperature was also important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ. The regression analysis reveals that mercury emission from soil (sediment)/air surface increased exponentially with raised air and soil (sediment) temperature. Hg concentration in the air was the most important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ at nighttime. Based on the measured data, we obtain regression equations. And we calculated the mean value of mercury exchange flux from soil (sediment)/air surface was12.77ng/m2h and annual Hg emission from soil (sediment) in WLFZ to the air to be19.53kg.
The mercury exchange fluxes on water/air surface were difference with water level fluctuated in different sampling periods. The highest value of mercury exchange fluxes on water/air surface was4.47±36.32ng/m2h (July, four times as great as other sampling periods), and the minimum value of mercury exchange fluxes on sediment/air surface was3.83±18.66ng/m2h (November). The biggest deposition was occurred in January. In general, the mercury exchange fluxes on water/air surface in WLFZ were similar to that of from ocean, estuary, reservoir and lake and lower than emission from large wetland systems at home and abroad. There were trend of diurnal variation characteristics of mercury exchange on water/air surface in WLFZ. At the daytime, a sharp increase in flux starting in the morning and peak emission at midday, then decreased gradually with sunlight decreased. At the nighttime, mercury exchange fluxes on water/air surface in WLFZ have been stabilized and its value remained at0ng/m2h fluctuation. Mercury exchange fluxes on water/air surface in WLFZ were dominated by emission at daytime in warm season and cold season. On the contrary, mercury exchange fluxes on water/air surface in WLFZ were dominated by deposition at daytime in cold season.
Solar radiation was the most important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ at daytime. The regression analysis reveals that mercury emission from water/air surface increased logarithmic (November, January and May) and power (July) with raised solar radiation. Temperature was also important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ. The regression analysis reveals that mercury emission from water/air surface increased exponentially with raised air and water temperature. Hg concentration in the air was the most important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ at nighttime. Based on the measured data, we obtain regression equations. And we calculated the mean value of mercury exchange flux from water/air surface was14.45ng/m2h and annual Hg emission from sediment in WLFZ to the air to be22.21kg.
The concentrations of total mercury (THg) in sediment in WLFZ were higher than the background value of sediment under the rivers and lower than that of sediment in other Hg polluted area. The concentrations of methyl mercury (MMHg) in sediment of WLFZ in July and September were0.128±0.028ng/g and0.031±0.027ng/g, respectively. MeHg:THg ratio in sediment of WLFZ in July and September were0.296±0.154%and0.069±0.081%, respectively. MeHg:THg ratio were difference with different type of sediment. And the results indicate that flooding sediment was bigger than semi-flooding sediment, and the value of drying sediment is least. In general, lower methylation potential was found in sediment of WLFZ.
The Hg contents in each fraction of sediments in July were different, Hg-e (46.99%) ranked first, followed by Hg-s (36.28%), Hg-o (9.57%), Hg-w (4.49%) and Hg-h (2.67%) in that order. In other words, Hg was mainly bound up with Fe and Mn oxides and amorphous organ-sulfur in July. The Hg contents in each fraction of sediments in September were also different, Hg-s (46.47%) ranked first, followed by Hg-e (34.73%), Hg-o (14.18%), Hg-h (2.40%) and Hg-w (2.21%) in that order. In other words, HgS was the main fraction of sediments in September.
With the flooding sediment exposed to the air, percentages of Hg in Hg-e of sediments were decreased and that of Hg-w and Hg-h were increased. It's called"spillover effects". With the waves and slope runoff bathed the drying sediment, Hg-s was preserved and Hg-w and Hg-h were taken away by erosion, causing HgS was becoming the main fraction of sediments. The existence of "spillover effects"was proved by the results of simulation testing in lab. The results of simulation testing in lab showed that with the water content in sediment decreased, percentages of Hg in Hg-e of sediments were decreased and that of Hg-w and Hg-h were increased. The mobility and bioavailability of Hg in flooding sediments were higher than that of in drying sediment. The process of flooding sediment exposed to the air, the mobility and bioavailability of Hg will increased. Meanwhile, when drying sediments were flooded, the mobility and bioavailability of Hg will also increase.
In the flooding period, MeHg:THg ratio in sediment of Treatment A (shallow sediment) and Treatment B (deepwater sediment) were0.41±0.29%%and0.74±0.52%, respectively. There existed statistically significant difference between the two methods with independent sample T test. In other words, shallow reservoir sediment has higher methylation potential than that of deepwater reservoir sediment. There existed no statistically significant difference among reactive mercury, dissolved mercury and methyl-mercury in overlying water with paired sample T test. There also existed no statistically significant difference among reactive mercury, dissolved mercury and methyl-mercury in pore water with paired sample T test. However, there existed statistically significant difference between overlying water and pore water for reactive mercury, dissolved mercury and methyl-mercury with paired sample T test.In other words, there existed gradients at the interfaces between overlying water and pore water for reactive mercury, dissolved mercury and methyl-mercury.
The diffusion flux of reactive mercury for Treatment A and Treatment B were12.79±5.06ng/m2d and13.24±3.68ng/m2d in different experimental flooding duration, respectively. The diffusion flux of dissolved mercury for Treatment A and Treatment B were154.65±47.12ng/m2d and160.23±56.19ng/m2d in different experimental flooding duration, respectively. The diffusion flux of methyl-mercury for Treatment A and Treatment B were7.61±3.39ng/m2d and7.79±4.56ng/m2d in different experimental flooding duration, respectively. There existed no statistically significant difference between Treatment A and Treatment B for the diffusion flux of reactive mercury, dissolved mercury and methyl-mercury with paired sample T test.
The contribution rate of reactive mercury in pore water to that of in overlying water was0.002-0.011%. The contribution rate of dissolved mercury in pore water to that of in overlying water was0.002-0.017%. And the contribution rate of methyl-mercury in pore water to that of in overlying water was0.015-0.143%.
水库是一个典型的“汞敏感生态系统”,水库环境汞环境化学行为的研究至关重要。三峡水库属特大型年调节水库,按照三峡水库“蓄清拍浊”的调度方式,水库正常蓄水位为175m,防洪限制水位为145m。这种水库调度方式使得水库周边形成垂直高度为30m、面积约350km2的水库消落带。一方面,由于是库区径流的汇集地带,消落带自然也成为环境汞的汇集区(汞汇)。另一方面,它所汇集的汞等污染物最终又会影响到水体质量(汞源)。同时,周期性的水位变化使得消落带将长期处于干湿交替的变换过程,消落带沉积物和上覆水体的性质将会发生一系列变化,如pH、Eh、电导率、微生物、有机质等,沉积物和水体环境的变化对汞的迁移转化以及水生生物的富集将有很大影响。
目前,对三峡库区汞问题的研究还仅限于对水体汞含量、鱼汞分布、淹没土壤汞含量的调查,而对消落带沉积物中汞的赋存形态转化规律与机制、界面释放特征及其生态效应等方面的研究还相当缺乏。基于此,本文采用野外现场调查采样分析与室内模拟实验研究相结合的方法,对三峡库区消落带沉积物汞在沉积物-大气界面、水体-大气界面、土壤(沉积物)水界面的释放特征,及沉积物中汞赋存形态、转化规律与机制进行了系统研究。
结果表明,在水位涨落的不同时期,库区消落带土壤(沉积物)-大气界面汞交换通量也是不同的。11月份土壤(沉积物)-大气界面汞交换通量最高,为28.17±36.17ng/m2h。1月份库区消落带土壤(沉积物)-大气界面汞交换通量最低,为-6.80±12.35ng/m2h,总体表现为沉降。总体来看,三峡库区消落带土壤(沉积物)-大气界面汞的交换通量高于国内外森林土壤和背景土壤汞的交换通量,但低于国内外对湿地沉积物、冲积平原土壤及农田土壤的释汞通量。三峡库区消落带土壤(沉积物)-大气界面汞的交换通量具有一定的日变化特征,晴天白天从早晨开始士壤(沉积物)-大气界面汞的交换通量逐渐增加,到正午左右达到峰值,之后交换通量逐渐减小。夜间土壤(沉积物)-大气界面汞交换通量的变化较为平稳,大部分数值在0ng/m2h附近波动。而阴天库区消落带土壤(沉积物)-大气界面汞交换通量变化的趋势不明显。暖季白天土壤(沉积物)-大气界面汞交换通量都以释放(正值)为主,冷季白天则以沉降(负值)为主。夜间,除11月份表现为释放外,其余三个采样时段夜间土壤(沉积物)-大气汞交换总体上均表现为沉降。光照强度是影响白天土壤(沉积物)-大气界面汞交换的最主要因素,随着光照强度的增强,土壤(沉积物)释汞通量呈线性(11月)、对数(1月)或乘幂(5月、7月)形式增加。温度也是影响库区消落带土壤(沉积物)-大气界面汞交换的重要因素,土壤(沉积物)温度和气温都与土壤(沉积物)-大气界面汞释放通量呈指数关系。大气汞浓度是影响消落带夜间土壤(沉积物)-大气界面汞交换的最主要因素。根据实测数据得到了基于诸影响因素的库区消落带土壤(沉积物)-大气汞交换通量的逐步回归方程,并据此估算了三峡库区消落带土壤(沉积物)-大气界面平均汞交换通量为12.77ng/m2h,库区消落带沉积物向大气释汞量为19.53kg/a。
库区消落带水位涨落的不同时期,水体-大气界面汞交换通量也是不同的。7月份水体-大气界面汞交换通量最高,为24.47±36.32ng/m2h,是其他采样时段水体-大气界面汞交换通量的约4-6倍。11月份库区消落带水体-大气界面汞交换通量最低,为3.83+18.66ng/m2h,最大沉降值(负值)还是出现在冷季的1月份。总体来看,三峡库区消落带水体-大气界面汞的释放通量与国内外学者对海洋、河口水域、水库、湖泊水体的释汞通量相近,但低于国外对大型湿地水体释汞通量的监测结果。三峡库区消落带水体-大气界面汞交换通量的具有一定的日变化特征,从早晨开始水体-大气界面汞的交换通量逐渐增加,到正午左右达到峰值,之后交换通量逐渐减小。夜间水体-大气界面汞交换通量的变化较为平稳,大部分数值在Ong/m2h附近波动。无论是暖季还是冷季,白天水体-大气界面汞交换通量都以释放(正值)为主。夜间,冷季库区消落带水体-大气界面汞交换通量以沉降(负值)为主,暖季界面汞的释放和沉降达到平衡。光照强度是影响白天水体-大气界面汞交换的最主要因素,随着光照强度的增强,水体释汞通量呈对数(11月、1月、5月)或乘幂(7月)形式增加。温度也是影响库区消落带水体-大气界面汞交换的重要因素,水温和气温都与水体-大气界面汞释放通量呈指数关系。大气汞浓度是影响消落带夜间水体-大气界面汞交换的最主要因素。根据实测数据得到了基于诸影响因素的库区消落带水体-大气汞交换通量的逐步回归方程,并据此估算了三峡库区消落带水体／大气平均汞交换通量为14.45ng/m2h,库区消落带水域向大气释汞量为22.21kg/a。
库区消落带沉积物总汞(THg)含量略高于河流沉积物汞的背景值,但比其他被汞污染的沉积物低很多,三峡库区消落带汞的污染程度较轻。7月份和9月份采集的库区消落带沉积物甲基汞(MeHg)的含量分别为0.128±0.028ng/g和0.031±0.027ng/g。7月份MeHg:THg的值为0.296±0.154%,9月份MeHg:THg的值为0.069±0.081%,两者相差近3倍。不同类型沉积物MeHg：THg的值也存在差异,总体表现为淹水沉积物>半淹水沉积物>落干沉积物。但总体来讲,三峡库区消落带沉积物中汞甲基化的速率和程度还处于较低的水平。7月份库区消落带沉积物中汞的赋存形态依次为强络合态(46.99%)、硫化物结合态(36.28%)、有机及其他络合态(9.57%)、水溶态(4.49%)、胃酸溶态(2.67%)。强络合态是7月份库区消落带沉积物中汞主要的赋存形态,即7月份库区消落带沉积物中汞主要是与铁锰氧化物、无定形有机硫化物等的络合物,以及元素汞和结合于矿物质晶格结构中的汞。9月份库区消落带沉积物中汞的赋存形态依次为硫化物结合态(46.47%)、强络合态(34.73%)、有机及其他络合态(14.18%)、胃酸溶态(2.40%)、水溶态(2.21%)。硫化物结合态是9月份库区消落带沉积物中汞主要的赋存形态,即9月份库区消落带沉积物中汞主要是与硫化物结合态的汞。伴随着淹水沉积物的出露,由于无机汞在相态分布中的溢出效应,强络合态的所占的比例逐渐下降,水溶态和胃酸溶态所占的比例逐渐升高。沉积物落干过程中,经历波浪和坡面径流对库岸沉积物的冲刷,导致弱结合态汞的流失,而强结合态汞如硫化物结合态的汞的比例的增加。室内模拟实验的结果也证实了沉积物在出露过程中,确实存在无机汞相态分布的“溢出效应”。即伴随着沉积物含水量的下降,强络合态的所占的比例逐渐下降,水溶态和胃酸溶态所占的比例逐渐升高。淹水沉积物中汞的迁移性要高于落干沉积物,沉积物出露过程中汞的迁移性增加,落干沉积物淹水后汞的迁移性也会增加。
淹水期间浅水沉积物(处理A)中甲基汞占总汞的比例为0.41+0.29%,深水沉积物(处理B)中甲基汞占总汞的比例为0.74±0.52%,两处理沉积物甲基汞占总汞的比例之间存在显著性差异(p=0.009),也就是说深水沉积物比浅水沉积物具有更高的甲基汞产率。处理A与处理B上覆水体中活性汞、溶解态汞、甲基汞含量之间没有显著性差异,沉积物孔隙水中活性汞、溶解态汞、甲基汞含量之间也没有显著性差异。但上覆水和沉积物孔隙水活性汞、溶解态汞、甲基汞含量存在显著的差异,也就是说活性汞,溶解态汞,甲基汞在沉积物孔隙水与上覆水之间存在浓度梯度,沉积物孔隙水中活性汞,溶解态汞,甲基汞有向上覆水中迁移扩散的趋势。不同淹水时间两处理沉积物孔隙水中活性汞的扩散通量分别为12.79±5.06ng/m2d(处理A)和13.24+3.68ng/m2d(处理B)；溶解态汞的扩散通量分别为154.65±47.12ng/m2d(处理A)和160.23±56.19ng/m2d(处理B)。甲基汞的扩散通量分别为7.61±3.39ng/m2d(处理A)和7.79±4.56ng/m2d(处理B)。各处理之间沉积物／水界面间活性汞,溶解态汞,甲基汞扩散通量没有显著性差异。库区消落带孔隙水中的活性汞对上覆水体活性汞含量的贡献率为0.002-0.011%,孔隙水中的溶解态汞对上覆水体溶解态汞含量的贡献率为0.002-0.017%,孔隙水中的甲基汞对上覆水体甲基汞含量的贡献率为0.015-0.143%。
Mercury and its compounds (especially alkyl mercury) have very strong biological toxicity. And the rate of biological enrichment and amplification were higher than other heavy metals. Meanwhile, their biological half-lifes in brain organ were also much longer. The problem caused by mercury pollution has been aroused public concern. The successful holding of10th International Conference on Mercury as a Global Pollutant (ICMGP) in Halifax, Nova Scotia, Canada in July24-29,2011. And the Intergovernmental Negotiation Committee (INC) organized by UNEP Ad Hoc Open Ended Working Group on Mercury has also been held three times since2010. All of these events do suggest that the focus of international research on mercury have been rise gradually in recent years.
The ecosystem of reservoir was very sensitive to mercury. It is an indisputable fact that the mercury concentration in fish will higher than other water body after the reservoirs were built. It has already become one of hot topics of mercury research. The largest reservoir in china, the famous Three Gorges Reservoir, is located at the middle reach of the Yangtze River. The water level of the Three Gorges Reservoir fluctuates from145m in summer (May-September) to175m in winter (October-April), and the water-level-fluctuation zone (WLFZ) with area of350km2were formed. On the one hand, WLFZ is an effective sink for mercury due to its integration and accumulation of mercury from both the terrestrial and aquatic environments (the surface water of adjacent river). And mercury from industrial activities and domestic consumption in the uplands may be accumulated in the WLFZ by flooding. On the other hand, the higher Hg concentration in sediment will also influence the quality of overlying water.
Periodic fluctuation of reservoir water caused the sediment could face a prolonged period of dry and wet alternate. And there will be a series of physical, chemical, mechanical changes (such as pH, Eh, electrical conductivity, microorganisms, organic matter etc.) in sediment and overlying water. There is, however, something that eventually will have a much bigger impact on Release, Movement and Transformation of Mercury in Water-Level-Fluctuating Zone of Three Gorges Reservoir Area, China.
Up to now, the research about mercury problem in WLFZ of Three Gorges Reservoir Area were just confined to Hg concentrations in water, the distribution of Hg in fish body and investigate the Hg concentrations in flooding soil. However, the speciation of mercury and its transformation in sediment in WLFZ, the characteristic of release from interface and its ecological effects were very insufficient. It's exactly because of this, our research were carried out based on field investigation, field sampling, and laboratory experiments. In this thesis, the author attempts to make a systematic analysis of mercury exchange fluxes on water/air, sediment/air and sediment/water interface in WLFZ. And the speciation of mercury and its transformation in sediment in WLFZ were also included.
The result showed that the mercury exchange fluxes on soil (sediment)/air surface were difference with water level fluctuated in different sampling periods. The highest value of mercury exchange fluxes on soil (sediment)/air surface was28.17±36.17ng/m2h (emission, November), and the minimum value of mercury exchange fluxes on sediment/air surface was-6.80±12.35ng/m2h (deposition, January). In general, the mercury exchange fluxes on soil (sediment)/air surface in WLFZ were higher than emission from forest soil and background soil at home and abroad, and lower than emission from sediment in wetland, soil in floodplain and soil in agriculture field. There were trend of diurnal variation characteristics of mercury exchange on soil (sediment)/air surface in WLFZ in sunny day. At the daytime, a sharp increase in flux starting in the morning and peak emission at midday, then decreased gradually with sunlight decreased. At the nighttime, mercury exchange fluxes on soil (sediment)/air surface in WLFZ have been stabilized and its value remained at0ng/m h fluctuation. There were non-significant trend of diurnal variation characteristics of mercury exchange on soil (sediment)/air surface in WLFZ in cloudy day. Mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by emission at daytime in warm season. On the contrary, mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by deposition at daytime in cold season. Except November, mercury exchange fluxes on soil (sediment)/air surface in WLFZ were dominated by deposition at nighttime.
Solar radiation was the most important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ at daytime. The regression analysis reveals that mercury emission from soil (sediment)/air surface increased line (November), logarithmic (January) and power (May and July) with raised solar radiation. Temperature was also important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ. The regression analysis reveals that mercury emission from soil (sediment)/air surface increased exponentially with raised air and soil (sediment) temperature. Hg concentration in the air was the most important factor which influenced the mercury exchange fluxes on soil (sediment)/air surface in WLFZ at nighttime. Based on the measured data, we obtain regression equations. And we calculated the mean value of mercury exchange flux from soil (sediment)/air surface was12.77ng/m2h and annual Hg emission from soil (sediment) in WLFZ to the air to be19.53kg.
The mercury exchange fluxes on water/air surface were difference with water level fluctuated in different sampling periods. The highest value of mercury exchange fluxes on water/air surface was4.47±36.32ng/m2h (July, four times as great as other sampling periods), and the minimum value of mercury exchange fluxes on sediment/air surface was3.83±18.66ng/m2h (November). The biggest deposition was occurred in January. In general, the mercury exchange fluxes on water/air surface in WLFZ were similar to that of from ocean, estuary, reservoir and lake and lower than emission from large wetland systems at home and abroad. There were trend of diurnal variation characteristics of mercury exchange on water/air surface in WLFZ. At the daytime, a sharp increase in flux starting in the morning and peak emission at midday, then decreased gradually with sunlight decreased. At the nighttime, mercury exchange fluxes on water/air surface in WLFZ have been stabilized and its value remained at0ng/m2h fluctuation. Mercury exchange fluxes on water/air surface in WLFZ were dominated by emission at daytime in warm season and cold season. On the contrary, mercury exchange fluxes on water/air surface in WLFZ were dominated by deposition at daytime in cold season.
Solar radiation was the most important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ at daytime. The regression analysis reveals that mercury emission from water/air surface increased logarithmic (November, January and May) and power (July) with raised solar radiation. Temperature was also important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ. The regression analysis reveals that mercury emission from water/air surface increased exponentially with raised air and water temperature. Hg concentration in the air was the most important factor which influenced the mercury exchange fluxes on water/air surface in WLFZ at nighttime. Based on the measured data, we obtain regression equations. And we calculated the mean value of mercury exchange flux from water/air surface was14.45ng/m2h and annual Hg emission from sediment in WLFZ to the air to be22.21kg.
The concentrations of total mercury (THg) in sediment in WLFZ were higher than the background value of sediment under the rivers and lower than that of sediment in other Hg polluted area. The concentrations of methyl mercury (MMHg) in sediment of WLFZ in July and September were0.128±0.028ng/g and0.031±0.027ng/g, respectively. MeHg:THg ratio in sediment of WLFZ in July and September were0.296±0.154%and0.069±0.081%, respectively. MeHg:THg ratio were difference with different type of sediment. And the results indicate that flooding sediment was bigger than semi-flooding sediment, and the value of drying sediment is least. In general, lower methylation potential was found in sediment of WLFZ.
The Hg contents in each fraction of sediments in July were different, Hg-e (46.99%) ranked first, followed by Hg-s (36.28%), Hg-o (9.57%), Hg-w (4.49%) and Hg-h (2.67%) in that order. In other words, Hg was mainly bound up with Fe and Mn oxides and amorphous organ-sulfur in July. The Hg contents in each fraction of sediments in September were also different, Hg-s (46.47%) ranked first, followed by Hg-e (34.73%), Hg-o (14.18%), Hg-h (2.40%) and Hg-w (2.21%) in that order. In other words, HgS was the main fraction of sediments in September.
With the flooding sediment exposed to the air, percentages of Hg in Hg-e of sediments were decreased and that of Hg-w and Hg-h were increased. It's called"spillover effects". With the waves and slope runoff bathed the drying sediment, Hg-s was preserved and Hg-w and Hg-h were taken away by erosion, causing HgS was becoming the main fraction of sediments. The existence of "spillover effects"was proved by the results of simulation testing in lab. The results of simulation testing in lab showed that with the water content in sediment decreased, percentages of Hg in Hg-e of sediments were decreased and that of Hg-w and Hg-h were increased. The mobility and bioavailability of Hg in flooding sediments were higher than that of in drying sediment. The process of flooding sediment exposed to the air, the mobility and bioavailability of Hg will increased. Meanwhile, when drying sediments were flooded, the mobility and bioavailability of Hg will also increase.
In the flooding period, MeHg:THg ratio in sediment of Treatment A (shallow sediment) and Treatment B (deepwater sediment) were0.41±0.29%%and0.74±0.52%, respectively. There existed statistically significant difference between the two methods with independent sample T test. In other words, shallow reservoir sediment has higher methylation potential than that of deepwater reservoir sediment. There existed no statistically significant difference among reactive mercury, dissolved mercury and methyl-mercury in overlying water with paired sample T test. There also existed no statistically significant difference among reactive mercury, dissolved mercury and methyl-mercury in pore water with paired sample T test. However, there existed statistically significant difference between overlying water and pore water for reactive mercury, dissolved mercury and methyl-mercury with paired sample T test.In other words, there existed gradients at the interfaces between overlying water and pore water for reactive mercury, dissolved mercury and methyl-mercury.
The diffusion flux of reactive mercury for Treatment A and Treatment B were12.79±5.06ng/m2d and13.24±3.68ng/m2d in different experimental flooding duration, respectively. The diffusion flux of dissolved mercury for Treatment A and Treatment B were154.65±47.12ng/m2d and160.23±56.19ng/m2d in different experimental flooding duration, respectively. The diffusion flux of methyl-mercury for Treatment A and Treatment B were7.61±3.39ng/m2d and7.79±4.56ng/m2d in different experimental flooding duration, respectively. There existed no statistically significant difference between Treatment A and Treatment B for the diffusion flux of reactive mercury, dissolved mercury and methyl-mercury with paired sample T test.
The contribution rate of reactive mercury in pore water to that of in overlying water was0.002-0.011%. The contribution rate of dissolved mercury in pore water to that of in overlying water was0.002-0.017%. And the contribution rate of methyl-mercury in pore water to that of in overlying water was0.015-0.143%.
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