A Review on the Impacts of Microplastics and Environmental Pollutants on Soil Microorganisms
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摘要:
随着塑料制品在工农业和生产生活中的广泛使用，大量微塑料被释放到土壤中，带来不可忽视的生态环境与健康风险。长久以来，人们更关注微塑料本身的生态毒性，对微塑料与环境中的其他化学污染物的联合作用及其环境效应研究较少。由于土壤微生物在微塑料降解过程中起着关键作用，认识土壤微塑料是如何通过影响土壤环境而直接或间接影响土壤中微生物群落和土壤生态功能的微观机理，已成为未来推进微塑料的降解和科学认识微塑料生态系统风险的关键。本文综述了近年来微塑料在土壤中吸附和迁移机理，以及微塑料的吸附程度和位置对其迁移行为的影响。总结了微塑料与土壤中有机污染物和重金属的复合作用的进展，探讨了这些复合作用对土壤环境风险的影响，包括污染物的毒性、生物利用度和迁移性的变化。评述了微塑料对土壤微生物群落的影响及作用机制，微塑料对微生物的物种丰富度、活性和结构的影响，以及微塑料表面的定殖和选择性富集能力。建议未来应该加强以下三方面的研究：①深入探索微塑料与环境污染物的复合作用及其生态毒理作用的微观机理；②认识土壤中微塑料对土壤微生物群落结构改变的微观机制；③探索通过科学调控土壤理化特性、特异性微生物在微塑料表面的定殖与富集能力等途径来控制土壤中微塑料及微塑料-其他环境污染物复合污染的可能性。
要点(1)微塑料在土壤中吸附迁移受到微塑料性质、土壤性质、动植物活动等多种因素的影响。
(2)微塑料与土壤中有机污染物和重金属的复合作用可能改变污染物的环境风险、迁移、降解和生物可利用度，这取决于微塑料和土壤的特性及环境因素。
(3)微塑料可增加或降低土壤微生物丰富度和多样性，该影响受土壤理化性质、微生物定殖和复合污染等多重机制调控。
HIGHLIGHTS(1) The adsorption and migration of microplastics in soil are influenced by various factors such as the properties of microplastics, soil, and flora and fauna activities.
(2) The combined effects of microplastics with organic pollutants and heavy metals in soil could alter the environmental risks, mobility, degradation, and bioavailability of pollutants. This alteration is contingent upon the characteristics of microplastics, soil, and environmental factors.
(3) Microplastics can either increase or decrease the richness and diversity of soil microorganisms, affecting them through multiple mechanisms regulated by soil physicochemical properties, microbial colonization, and composite pollution.
Abstract:The extensive use of plastic results in a significant release of microplastics into the soil, posing risks to ecosystems and human health. Research on the interaction between microplastics and pollutants and their combined effects is sparse. Understanding how soil microplastics affect microbial communities is crucial for assessing ecological risks. This comprehensive review examines the adsorption and migration mechanisms of microplastics, with a specific focus on their impact on migration. It explores the combined effects of microplastics with organic pollutants and heavy metals, leading to changes in toxicity, bioavailability, and mobility. Additionally, the review investigates how microplastics influence soil microbial communities, revealing alterations in species richness, activity, and structure. The findings of this review highlight the significant impact of microplastics on pollutants, the modifications in toxicity, bioavailability, and mobility of combined pollutants, as well as their influence on soil microbial communities. To comprehensively assess the environmental impact, it is essential to understand how microplastics interact with pollutants. The review underscores the need to comprehend their influence on soil microbes and functions in order to effectively address ecological risks. Future research should prioritize exploring microscale mechanisms and developing strategies to mitigate soil microplastics and associated pollution. The BRIEF REPORT is available for this paper at http://www.ykcs.ac.cn/en/article/doi/10.15898/j.ykcs.202209180175.
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Keywords:
- microplastics /
- microbial community structure /
- organic pollutants /
- heavy metals /
- synergistic effects
BRIEF REPORTUnderstanding the impact of microplastics (MPs) is crucial due to their persistent presence in the environment and global consequences[1-2]. Ranging from 100 to 5mm, these pollutants undergo processes that can break them down into smaller sizes (<100nm) through mechanisms like photodegradation or environmental wear. The intricate ecological effects of microplastics are evident in diverse environmental media, including soil, water, and air[4-6]. The urgency for investigation is highlighted by the prevalence of microplastics in remote areas like the Arctic and their identification in human blood, urine, and feces[7-12]. The environmental impact of microplastics extends to potentially accelerating Arctic ice melting[9], posing significant health risks to multiple human organ systems[13-14]. Concerns also arise from harmful additives released by microplastics, potentially disrupting normal human growth and development[16]. In soil environments, microplastics display adsorption and migration behaviors[17-18], acting as carriers that influence the mobility and availability of organic pollutants and heavy metals[19-21]. However, a detailed understanding of the interaction mechanisms between microplastics and pollutants requires further exploration. Another critical research area involves investigating the impact of microplastics on soil microbial communities, with recent studies indicating alterations in microbial diversity, biomass, and functional gene expression due to microplastic exposure.
Addressing scientific gaps is imperative for a holistic understanding. Exploring the adsorption and migration mechanisms of microplastics under various soil types and environmental conditions, comprehensively deciphering the synergistic effects between microplastics and organic pollutants or heavy metals, and unraveling the specific mechanisms and ecological implications of microplastics on soil microorganisms are essential. This comprehensive review aims to shed light on the characteristics and microscopic mechanisms of microplastics, advancing our comprehension of their implications for soil health and ecology.
1. Adsorption and migration of microplastics in a soil environment
Microplastics interact with soil primarily through electrostatic forces and physical retention[22]. Their surface charge influences adsorption and retention, affecting interactions with environmental ions. For example, when adsorbing onto kaolin clay[23], hydrophobic interactions and hydrogen bonding dominate, especially with polyamide’s polar amide groups. Initially positively charged (MPs+)[24], microplastics weather to develop a negative charge (MPs−)[25]. Changes in soil conditions may alter soil adsorption capacity, potentially enabling microplastic migration[26]. The migratory behavior varies among different types of microplastics. Fibrous microplastics exhibit stronger migration compared to agricultural film fragments and fragmented microplastics[27]. Additionally, low-density microplastics are more susceptible to lateral migration in soil or water, influenced by natural forces[28]. Environmental changes in soil conditions[30], disturbances caused by flora and fauna[16,31-32], and anthropogenic activities[27] collectively influence the intricate process of microplastic migration in soil.
2. Interaction of microplastics with organic contaminants and heavy metals in soil The complex interplay between microplastics, organic pollutants, and heavy metals in soil has generated ongoing debates regarding its impact on environmental risk. Microplastics, such as polyethylene (PE) and polypropylene (PP), have been shown to adsorb organic contaminants, potentially reducing the free fraction of these pollutants and mitigating soil toxicity[36-37]. However, microplastics can also increase the overall residue levels of organic contaminants in soil, exemplified by a rise in pesticide residues from 4% to 15% due to microplastic presence[39]. The effectiveness of microplastics in adsorbing organic pollutants is influenced by soil factors. With a high specific surface area and hydrophobic properties, microplastics efficiently carry hydrophobic organic compounds (HOCs)[40]. Soil organic matter (SOM) also affects their adsorption capacity, indirectly influencing the distribution and bioavailability of soil polycyclic aromatic hydrocarbons (PAHs)[42]. Additionally, microplastics’ structural composition and aging impact their adsorption abilities, showing varied capacity among different types[45-47]. Environmental factors like temperature, pH, salinity, and ion strength further modulate organic compound adsorption by microplastics[47-48]. Their impact on contaminant degradation relies on complex microorganism, pollutant, and microplastic surface interactions[49].
In the context of the interaction between microplastics and heavy metals in soil, studies have shown that joint treatment with 0.1% PE-microplastics and heavy metals can lead to increased accumulation of certain heavy metals in plants, such as Cu and Pb in rapeseed[53]. Additionally, PE-microplastics can enhance the bioavailability of Cd in soil, affecting its accumulation in lettuce[54]. However, it is important to note that microplastics may physically damage plants, exacerbating the toxicity of heavy metals[58]. Nevertheless, some studies have not found significant effects of microplastics on the absorption of heavy metals by plants[59-60].
The adsorption-desorption mechanisms between microplastics and heavy metals in soil are complex, with microplastics capable of adsorbing various heavy metals from their surroundings[61-69]. Factors like surface charge, pH, organic matter on microplastic surfaces, and the type of microplastic all influence this process[66-71]. Additionally, the competitive adsorption of heavy metal ions on microplastic surfaces can lead to partial desorption of some heavy metals[54].
In summary, the interaction of microplastics with organic pollutants and heavy metals in soil is multifaceted, influenced by various environmental factors and the specific properties of microplastics. Understanding these interactions is crucial for assessing their impact on soil contamination and ecosystem health.
3. Impacts and mechanisms of microplastics on soil microbial communities
Microplastics have emerged as influential agents affecting soil microbial communities, with profound implications for biodiversity and ecosystem functions.
Enhancement effects: Microplastics have been observed to augment the abundance and diversity of specific microbial communities[10,78]. Notably, genera like Pseudomonas and Nitrospira exhibit increased abundance in the presence of microplastics[79-80]. Within microplastic-treated soils, a significant increase in the gene abundance of Nitrospira, crucial for nitrification, results in a reduction in NH4+-N content[81]. Polyethylene (PE) and polyvinyl chloride (PVC) microplastics promote the proliferation of microbial communities associated with membrane transport functions[51,82].
Detrimental effects: Contrarily, studies have also highlighted adverse impacts of microplastics on soil microbial communities. For instance, residual film microplastics in agricultural soils intensify bacterial community succession, destabilizing the microbial community structure and compromising soil functions[86]. Specific microplastic types, like polystyrene nanoparticles (PS-NPs) and certain polyethylene microparticles (PE-MPs), have shown to significantly alter fungal community compositions, with fungi being more sensitive to microplastic presence than bacteria[10,87-88].
Mechanistic insights: The interaction between microplastics and soil microbial communities is underpinned by alterations in soil physicochemical properties, such as aggregation, bulk density, and nutrient status, directly influencing microbial colonization and enrichment[90-92]. Soil biofilm-associated microbial communities exhibit marked differences in composition and genetics compared to adjacent soil environments[95-96]. This suggests that microplastics may selectively enrich specific microbial taxa. Furthermore, microorganisms have evolved diverse adsorption mechanisms to adapt to the presence of microplastics, underscoring the dynamic nature of soil-microplastic interactions[49].
Concluding remarks: While certain studies report no significant impacts of microplastics on soil bacterial communities under specific conditions[89], the overarching consensus underscores the intricate interplay between microplastics and soil microbial dynamics, necessitating further research to elucidate long-term ecological consequences.
This streamlined overview encapsulates the multifaceted relationship between microplastics and soil microbial communities, emphasizing both their beneficial and detrimental impacts, while highlighting the need for continued academic exploration.
4. Future perspectives
Future research should prioritize: (1) Investigating the impact of microplastic co-pollution with contaminants (organic pollutants, heavy metals) on soil microbial communities. (2) Exploring microbial pathways for safe microplastic degradation, with potential applications for environmentally sound removal. (3) Developing strategies to improve soil microbial community structures by regulating physicochemical properties, specifically focusing on microbe colonization and enrichment on microplastic surfaces. This aims to facilitate safer regulation and degradation of microplastics, particularly in scenarios of compound pollution with other contaminants.
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西藏处于典型的喜马拉雅地热带，是中国高温地热流体分布最密集的地区，其地热资源居中国第一［1］，境内共有709个地热带活动区，其中有131个地热系统温度高于150℃，8处地热层温度高于200℃［2］。位于境内的羊八井地热站是中国最大的地热发电站，也是中国所有热液系统中测得的储层温度最高的地热田［3］，其日开采汽水总量约为12000t［2］。砷(As)和氟(F)是西藏羊八井高温地热流体中两种典型的高浓度有害元素，通过地热开发可以进一步促进或加速地热源As和F向地表或近地表环境释放，从而威胁附近水土生态环境。因此，调查As和F浓度水平与水体水化学特征从而揭示水体As和F的富集规律，对丰富和认识西藏地区水环境中As和F的环境地球化学行为具有重要意义。
关于羊八井地热水的水化学特征、水热蚀变和水体胶体粒子特征已开展前期研究［4-6］。研究表明，羊八井地热田热水水化学类型为Na-Cl型，大部分地热水为中性偏碱性，pH值在6.70～9.60之间，浅层热水主要来源于深层热水与地表水(冰川雪山融水和大气降水)的混合［7-8］。地表水补给来源于念青唐古拉山海拔4400～5800m的雪山融水［8］，融雪水渗入地下后由底层岩浆热源加热，由于热水密度低于融雪水，加热后的热水能自然地流向地表。相比水化学研究，羊八井地热储层流体中As和F浓度特征、地热源As和F引起的水土环境影响是地热利用过程中关注的热点环境问题［9-12］。郭清海等［13］报道了羊八井热田地热流体As和F浓度分别高达5.70mg/L和19.60mg/L，远高于西藏其他地区［14-15］，由于地热开发，输入地热邻近河流堆龙曲中主要的污染物为As和F。魏晓阳等［11］研究表明地热邻近河流堆龙曲中检出了高浓度F(0.41～1.31mg/L)。在As和F来源及富集机制方面，氟化物浓度受到氟石(CaF2)溶解度限制，其浓度与水化学类型密切相关，表现为F元素在Na-Cl或Na-Cl·SO4等Na型水中富集程度常高于Na·Ca-HCO3或Ca-HCO3等Ca型水体［15］，因此羊八井高温地热水氟化物浓度高达19.60mg/L［13］。同时，弱碱性pH水体也为As和F的富集提供了有利条件［16］。羊八井地热流体中As主要来自岩浆脱气［17］，As的富集与岩浆流体的浸取和地幔侵入高砷岩浆热源有关［18］，也就是说，决定岩浆热液流体中As浓度高低最关键的因素是岩浆流体的地质成因及其化学成分。综上所述，羊八井高温地热水中As和F的来源及富集机制主要有两方面：①深部地热流体的升流混合作用；②补给水向下渗流过程中与含As和F硅酸盐矿物的溶滤作用，因此地下水体中As和F往往表现为共生性［19-20］。尽管研究者在羊八井地热水中As和F的浓度特征及其来源方面有了前期研究基础，但仍需要深入分析地热源As和F浓度年变化趋势、水环境演变规律和水土环境生态风险。
本文在前期研究的基础上，于2021—2022年对西藏羊八井地热田进行了三期的地热水、温泉水和土壤样品采集，分析羊八井热田水化学常规理化指标、水化学类型、阴阳离子组成，揭示地热水、温泉水和土壤样品中As和F浓度变化特征，剖析地热水和温泉水As和F的来源与富集机制，评价了水体和地表土壤超标风险情况，研究成果为羊八井地区地热的持续合理利用提供科学依据。
1. 研究区域地理位置概况
西藏羊八井地热发电站位于西藏自治区拉萨市西北约90km的当雄县羊八井镇，海拔约4300m。羊八井地层主要由第四系沉积物和基岩风化壳组成，第四系沉积物主要为冲洪积砂砾石层和冰碛砂砾层，而基岩风化壳则主要由花岗岩风化而成。羊八井气候寒冷干燥，年平均气温在2.5℃左右，年降水量在500mm左右［2］。由于空气稀薄，太阳辐射强，日照时间长，全年无霜期短，羊八井热田是中国目前已知的热储温度最高的地热田，其深部热储平均温度为252℃，最高记录热储温度达329.8℃［2］，地表出露温度为68～84℃［21］。羊八井地热发电厂是中国建设的第一座最大的、海拔最高的地热试验田，也是当今世界迄今为止唯一利用中温浅层热储资源发电的电厂。地热站的建设为西藏地区的经济发展和社会稳定发挥了举足轻重的作用。羊八井地热站地势平坦，海拔7000m以上的念青唐古拉山屹立于地热站西北，东南方向为海拔6000m以上的唐山，终年覆盖有大量冰川，是地表径流的重要补给，地势上具有西北高、东南低的特点［22］。羊八井地热发电站和周边地理分布格局如图1所示。
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图 1 羊八井地热发电站位置(a)及周围分布格局(b)Figure 1. Geographical location (a) and surrounding distribution pattern (b) of Yangbajing geothermal power plant青藏公路将热田分为南北两区，北区分布有二电厂和国家地质公园，南区分布有一电厂，一电厂紧邻藏布曲。羊八井镇因地热资源丰富而闻名，镇内分布有规模宏大的喷泉、沸泉、涌射泉、热泉和热水湖等。在羊八井镇格达乡建设有规模仅次于羊八井地热站的羊易地热站。羊八井地热温泉洗浴已成为重要的旅游胜地，目前建成“蓝色天国”温泉旅游区。地热站于2019年进入休采期，休采期间“蓝色天国”温泉旅游区对外开放，钻井口仍有地热水流出，在地热休采期间，其水环境影响仍不容忽视。
2. 实验部分
2.1 样品采集
根据资料显示和现场勘察，三次采样分别于2021年6月(丰水期)、2021年11月(平水期)和2022年4月(枯水期)在羊八井地热发电站钻井口、温泉水口共设2个采样点，水样采集依照《水和废水监测分析方法》(GB/T 8538—2008)进行。现场采集2份平行水样约1L，测定常规指标，包括pH值、电导率(EC)、总溶解固体(TDS)和盐度(SAL)。测定方法是将优特PCS Testr 35型便携式多参数测量仪电极深入水面下10cm处［23］，待显示数字稳定后进行读数记录。地热水和温泉水出露温度采用水银温度计现场测定。
土壤样品全部采集于温泉排废口，使用铁锹采集温泉水淋滤的土壤约2kg，沥水冷却后装于塑料密封袋中保存，土壤样品带回实验室自然风干，过100目筛保存备用。水样带回实验室自然冷却，一份经0.45μm滤膜过滤后，加入优级纯硝酸5mL酸化并保存在0～4℃冰箱中备用，用于As和其他元素分析；另一份水样过滤后用于F离子和其他阴离子分析。
2.2 分析仪器
原子荧光光谱仪(AFS-9330型，AFS-8300型，北京吉天仪器有限公司)、高精度X射线荧光光谱仪(HD Rocksand型，美国XOS公司)、离子选择性电极(F090 ION 700型，美国Thermo Eutech公司)、电感耦合等离子体发射光谱仪(Optima 5300 DV型，美国PerkinElmer公司)、电感耦合等离子体质谱仪(ELAN DRC-e型，美国PerkinElmer公司)和离子色谱仪(IC，ICS-1000型，美国Dionex公司)用于目标物测定。
2.3 样品分析方法
(1) 水体和土壤中总As浓度采用原子荧光光谱法(AFS)和高精度X射线荧光光谱法(XRF)测定。水体总As浓度测定方法：地热水和温泉水逐级稀释200倍后，在9mL稀释后的待测水样中加入1mL 5%(硫脲+抗坏血酸)溶液，30℃恒温水浴反应30min，标准曲线的不同浓度点采用上述相同的操作进行，反应结束后进行总As浓度测定。土壤总As浓度测定方法：XRF测定土壤总As时，先用仪器自带能量校准样品(A750)进行能量校正，使用标准品(GBW07310)对仪器主要参数进行实验调试，以消除或减少元素间干扰，提高仪器准确度。
(2) F离子浓度采用离子选择性电极法测定。土壤中总F浓度的测定方法：依据《土壤质量 氟化物的测定 离子选择电极法》(GB/T 22104—2008)。具体流程：称取0.20g土壤样品于坩埚中，加入2g氢氧化钠，高温550℃条件下熔融煅烧，煅烧后采用热水浸取并定容至100mL，测定前加入适量盐酸中和到pH为5～6，采用氟电极测定F离子浓度。10mL样品中加入1mL总离子强度调节缓冲溶液(TISAB)并以掩蔽溶液中Fe3+和Al3+干扰。水体中总F浓度的测定与土壤中F的浸取液测定方法相同。
(3) 水体中元素Ca、K、Na、Mg、Fe、Al和Mn采用电感耦合等离子体发射光谱法(ICP-OES) 测定；Zn、Cr、Co、Ni、Mn、Cu和Cd 等元素采用电感耦合等离子体质谱法(ICP-MS) 测定，用浓度为10.00μg/L的Ba、Be、Ce、Co、In的调谐液优化仪器检测条件，使仪器灵敏度、氧化物离子产率、双电荷离子产率等各项指标达到测定要求。ICP-MS/OES元素分析采用在线加入内标物(In/Rh) 的方法降低基体干扰。水体Se、Hg和Sb通过原子荧光光谱仪测定。地热水和温泉水中阳离子(K+、Na+、Ca2+和Mg2+)和阴离子(Cl−和NO3 − )采用离子色谱法(IC)测定，CO3 2−和HCO3 − 采用容量法测定。水体阴阳离子IC分析和元素ICP-MS/OES分析是委托具有权威资质的第三方测试平台(西藏自治区地质矿产勘查开发局中心实验室)完成，样品测定值均为3次平行测定的平均值扣除空白后的结果，标准偏差小于5%。
目标物的分析方法和测定条件及检出限如表1所示。
表 1 样品分析方法及测定条件Table 1. Sample analysis methods and measurement conditions样品类型和元素 分析方法 检出限 RSD 仪器测定条件 水体As、Hg、Sb、Se AFS As：0.0096μg/L
Hg：0.0017μg/L
Sb：0.01μg/L
Se：0.01μg/L＜5% (1)还原剂：0.5% (m/m) NaOH+2% (m/m) KBH4
(2)载液：5% (V/V)盐酸
(3)载气(Ar)流速0.4L/min土壤As XRF 1mg/kg ＜5% 分析线Kβ；能量11.72keV；电压50kV；分析时间300s；滤光片Ag 土壤和水体F ISE 定量下限0.09mg/L ＜5% 10mL样品+1mL总离子强度调节缓冲溶液(TISAB) 水体Ca2+、K+、Na+、Mg2+、Cl−、NO3 − IC Ca2+：0.011mg/L
K+：0.02mg/L
Na+：0.005mg/L
Mg2+：0.013mg/L
Cl−：0.032mg/L
NO3 −：0.054mg/L＜5% (1) EGC-III淋洗液自动发生器；DS6型电导检测器
阳离子测定条件：CSRS 300-4 mm阳离子抑制器；CS12A型分离柱(4mm×250mm)；淋洗液20mmol/L硫酸；流速1mL/min；进样体积500μL
(2)阴离子测定条件：ASRS 300-4 mm阴离子抑制器；Ion Pac AS19型分离柱(4mm×250mm)；淋洗液：30mmol/L KOH; 流速1mL/min；进样体积500μL水体CO3 2−、HCO3 − 容量法 － ＜1% 5%酚酞-乙醇指示剂；1%溴酚蓝指示剂；双指示剂滴定分析法 水体Ca、K、Na、Mg、Fe、Al、Mn ICP-OES Ca：0.003mg/L
K：0.06mg/L
Na：0.02mg/L
Mg：0.02mg/L
Fe：0.002mg/L
Al：0.03mg/L
Mn：0.005mg/L＜5% (1)射频功率1250W；等离子体气(Ar）流速15L/min；辅助气(Ar）流速0.2L/min；雾化器气体(Ar）流速0.75L/min；样品提升量1.5L/min；观测方式：垂直；冲洗时间30s；积分时间5s；重复测定3次
(2)最佳波长选择：Ca 317.933nm、K 766.49nm、Na 588.995nm、Mg 285.213nm、Fe 238.204nm、Al 396.153nm、Mn 285.213nm水体V、Be、Zn、Cr、Co、Ni、Mn、Pb、Mo、Ti、Cu、Ba、Cd ICP-MS
Zn、Cr、Be、Co、Ni、Mn、Cu、Cd： 1～10ng/L；
Mo、Pb、Ba、Ti、V： 0.1～1ng/L＜5% (1)射频功率1150W；等离子体气(Ar）流速17 L/min；辅助气(Ar）流速1.2 L/min；载气(Ar）流速1.06 L/min；扫描模式为跳峰；重复测定3次
(2)m/z: 51V、9Be、66Zn、52Cr、59Co、60Ni、55Mn、208Pb、98Mo、48Ti、63Cu、115Ba、111Cd2.4 测试数据质量控制
水体总As和F浓度测定结果采用加标回收率的方法进行了准确性验证，结果如表2所示，水体总As的回收率在103.00%～114.80%，F的加标回收率在98.20%～99.90%。F的测定较As更准确，主要是F离子选择性电极法测定浓度为mg/L水平，而原子荧光光谱法测定总As浓度在μg/L水平，因而F的测定准确度更高。总体而言，As和F的加标回收率结果都在理论范围(80%～120%)，表明测定方法可靠。在土壤总As和F含量测定中，采用国家一级标准物质的方法对分析方法准确度进行了检验。选择了沉积物标准品GBW07310作为分析样品，在相同分析方法下进行测定，总As和F的回收率分别为113.5%和92.6%，总As和F的测定值与标准值吻合，综上所述，方法的准确性良好，数据可靠。
表 2 水体和土壤中As和F浓度测定准确性验证Table 2. The accuracy of measuring As and F concentrations in water and soil samples样品类型 元素 加标值
(mg/L)测定值
(mg/L)回收率
(%)地热水 As 0 3.16±0.10 − 3 6.25±0.12 103.0±2.69 F 0 15.91±0.24 − 15 30.65±0.47 98.2±3.16 温泉水 As 0 4.18±0.07 − 4 8.78±0.12 114.8±2.98 F 0 17.67±0.23 − 20 37.65±0.15 99.9±0.75 土壤样品 元素 标准值
(mg/kg)测定值
(mg/kg)回收率
(%)沉积物GBW07310 As 25±3.0 28.4±0.85 113.5±3.40 F 149±25 138.0±16.57 92.6±11.10 3. 结果与讨论
3.1 水体水化学类型
地热水主要用于电热发电厂发电，而温泉水主要用于蓝色天国洗浴中心。水质常规理化参数如表3所示，钻井口地热水出露温度在76～78℃之间，出露温度随季节性变化差异较小。温泉水水温在28.30～41.40℃，温泉洗浴水入口温度为41.40℃，温泉利用后，随着冷却水和生活水的共排放，温度会降低，排废温度在28.30～29.60℃。温泉水pH范围在7.87～9.42之间，入口pH值更高，排废口pH值低，地热水pH在8.95～9.15之间。温泉水和地热水的电导变化范围在1670～1882μS/cm之间，TDS值在1126～1340mg/L之间，盐度在914～983mg/L之间，水质变化基本呈现枯水期＞平水期＞丰水期的趋势。
表 3 水质常规理化参数Table 3. Conventional physicochemical parameters of the water quality采样时间 水期 样品类型 采样位置 水温
(℃)pH 电导
(μS/cm)TDS
(mg/L)盐度
(mg/L)2021年6月 丰水期 温泉水 温泉洗浴入口 41.4 9.42 1690 1180 914 地热水 电站钻井口 76.0 9.15 1699 1220 952 2021年11月 平水期 温泉水 温泉洗浴排废口 29.6 7.87 1882 1340 956 地热水 电站钻井口 78.0 8.95 1670 1213 935 2022年4月 枯水期 温泉水 温泉洗浴排废口 28.3 7.93 1783 1238 974 地热水 电站钻井口 77.5 9.14 1678 1126 983 以平水期地热水和温泉水样品为代表，测定了8个阴阳离子浓度，结果如表4所示，水体中阳离子Na+占主导，温泉水中的阴离子HCO3 − 和SO4 2−占主导，而地热水中阴离子Cl− 和HCO3 − 占主导。此外，阴阳离子平衡和相对误差也列入表4中，地热水阴阳离子平衡相对误差小于5%，表明分析数据可靠。在温泉水中，阴阳离子平衡相对误差较高，最高达13.15%，因为温泉水Na+占比较高，Cl−占比太低，导致阴阳离子平衡失调，从而导致相对误差偏高。
表 4 地热水和温泉水中主要阴阳离子浓度Table 4. Concentrations of major anion and cation ions in the geothermal and hot spring waters样品类型 阳离子浓度(mg/L) 阴离子浓度(mg/L) 阳离子当量浓度
(mmol/L)阴离子当量浓度
(mmol/L)相对误差
(%)Ca2+ Mg2+ K+ Na+ Cl− SO4 2− CO3 2− HCO3 − 地热水
(钻井口)一电厂6.66 0.20 35.19 445.5 331.6 16.79 85.95 546.4 20.63 21.53 2.13 地热水
(废井口)二电厂3.31 0.013 3.42 147.2 59.26 26.30 14.51 260.1 6.66 6.97 2.28 温泉水洗浴
(入口)38.23 10.86 15.91 183.4 72.46 99.9 4.84 262.3 11.19 8.59 13.15 温泉水洗浴
(排废)31.45 5.66 47.02 328.5 91.60 229 ND 428.9 17.53 14.38 9.85 注：“ND”为未检出。 采用Origin 9.2软件绘制了水体水化学Piper三线图，如图2所示，Piper三线图左侧三角形体现了主要阳离子的比例，右侧三角形体现了主要阴离子的比例，中间菱形体现了主要阴阳离子情况。数据点往高占比区域分布，表明水化学类型主要为高占比区域类型。地热水数据点集中分布在低Ca2+、高K++Na+占比方向，占阳离子总量的80%以上，而地热水中Na+浓度占K++Na+总浓度的87.50%～97.70%，因此阳离子主要以Na+为主导。水体Na+浓度高达445.5mg/L，Ca2+浓度低至3.31mg/L。在右下角的主要阴离子分布图中，阴离子主要分布在低Cl−和高CO2− 3+HCO− 3占比方向，而地热水中HCO− 3的浓度占CO2− 3+HCO− 3总浓度的86.50%～100%，因此阴离子主要以HCO− 3为主导。此外地热水Cl−占比最高达60%，综上所述，地热水水化学类型为Na-HCO3∙Cl，与文献[10]报道一致。
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图 2 水体水化学成分Piper三线图Figure 2. Piper three-line diagram of the hydrochemical composition of water samples3.2 水体砷和氟浓度
高温地热水中As的来源主要为岩浆脱气，表现为深层地热水中总As浓度(5.70mg/L)大于浅层地热水(2.99mg/L)［24］。此外，As的浓度还会受到水体pH值影响，在碱性条件下硫代砷酸盐浓度占比高达83%［25］。羊八井地热水和温泉水总As浓度如图3a所示，钻井口地热水浓度在3.16～3.56mg/L之间，平均值为3.32mg/L，几乎不随水期的变化而变化。地热水总As浓度与张庆等［10］报道的羊八井地热水中总As浓度(3.54～3.56mg/L)一致。同时，对比了位于同一流域上游的羊易电站地热水中总As浓度，羊八井地热水中总As浓度高于羊易电站(2.24～2.30mg/L)，这主要是地热区岩浆背景不同。相比钻井口地热水，温泉水总As浓度在4.18～6.50mg/L，浓度更高。地热水中F离子浓度如图3b所示，羊八井钻井口地热水F离子浓度在15.90～16.20mg/L，几乎不随水期的变化而变化，远高于西藏日多温泉中F离子浓度(6.20mg/L)。同时，对比了同一流域上游的羊易电站地热水中F离子浓度，羊八井地热水中F离子浓度显著(P＜0.01)高于羊易电站钻井口(12.63mg/L)和羊易电站喷射泉(5.19mg/L)。显著(P＜0.01)高于其他地区地热水(1.00～12.70mg/L)［26-27］，与文献[28-29]报道的F离子浓度(17.00～18.90mg/L)相当，低于美国黄石公园报道的热泉氟化物浓度(31.60mg/L)［30］。地热水总F离子浓度与2012年报道的羊八井地热水(18.00～18.90mg/L)和羊易电站地热水(19.20mg/L)［24］相比，F离子浓度呈下降趋势。与As的分布规律一致，温泉洗浴水中总F离子浓度更高，在14.56～17.89mg/L，这是因为受水气蒸发浓缩影响(图3c)，温泉水中As和F的浓度高于地热水。地热水来源于钻井口，水汽未发生分离或分离较少(图3d)，因而浓度低于温泉水。温泉水中As和F浓度呈现枯水期＞平水期＞丰水期，这与季节蒸发量大小一致。地热水和温泉水As和F浓度显著高于《地热资源评价方法》(DZ40—85)对有害成分规定的最高允许排放浓度(总As为0.50mg/L，氟化物为10mg/L)，地热废水的不当处理存在As和F的暴露风险。
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图 3 水体中总As(a)和F浓度(b)，温泉水水汽蒸发(c)和地热钻井口(d)照片Figure 3. The concentrations of total As (a) and F (b) in water samples, and pictures of hot spring (c) and geothermal water (d)As和F是高温地热水中典型的高浓度有害元素，主要来源于深层岩浆。通常，一些羟基矿物如白云母和黑云母常与F离子发生离子交换，当地下水为碱性时交换作用更容易发生，OH−能取代含F矿物质中的F离子，增加了地热水中F离子浓度，其基本过程存在如下反应［31］：
白云母：
$$ \begin{split} & \mathrm{K}\mathrm{A}\mathrm{l}_2\left[\mathrm{A}\mathrm{l}\mathrm{S}\mathrm{i}_3\mathrm{O}_{10}\right]\mathrm{F}_2+2\mathrm{O}\mathrm{H}^-= \\ &\mathrm{\ \ \ K}\mathrm{A}\mathrm{l}_2\left[\mathrm{A}\mathrm{l}\mathrm{S}\mathrm{i}_3\mathrm{O}_{10}\right]\left[\mathrm{O}\mathrm{H}\right]_2+2\mathrm{F}^- \end{split} $$ (1) 黑云母：
$$ \begin{split} & \mathrm{K}\mathrm{M}\mathrm{g}_3\left[\mathrm{A}\mathrm{l}\mathrm{S}\mathrm{i}_3\mathrm{O}_{10}\right]\mathrm{F}_2+2\mathrm{O}\mathrm{H}^-= \\ &\mathrm{\ \ \ K}\mathrm{M}\mathrm{g}\left[\mathrm{A}\mathrm{l}\mathrm{S}\mathrm{i}_3\mathrm{O}_{10}\right]\left[\mathrm{O}\mathrm{H}\right]_2+2\mathrm{F}^- \end{split} $$ (2) 羊八井浅层地热流体，pH值为8.95～9.15，偏碱性的水体为As和F溶出提供了有利条件。此外，水体中F离子浓度受氟石(CaF2)溶解度限制，Ca2+浓度越低，溶液中F离子浓度就会越高，而地热水中Ca2+浓度低至3.31mg/L，低浓度Ca2+是地热流体中F离子浓度富集的另一有利条件。
3.3 土壤砷和氟浓度
图4a为温泉排废口废水淋滤的土壤样品总As浓度。如图所示，枯水期和平水期总As浓度变化差异不大，总As浓度在97.60～99.08mg/kg之间，显著(P＜0.01)大于丰水期浓度(79.50mg/kg)。丰水期土壤浓度较低，可能是河水受雨水补给，周边泥沙冲刷，稀释了土壤中总As的量，使其浓度偏低。土壤中总As浓度随季节性变化差异不大。与西藏土壤总As背景值(18.70mg/kg)相比［32］，地热区的总As浓度显著高于背景值(P＜0.01)，是背景值的4.25～5.31倍，表明地热区土壤浓度处于高As污染风险。与济南温泉水尾水土壤中As浓度相比(15.45μg/kg)［33］，羊八井地热区As浓度约高出3个数量级。与污染的寨上金矿矿区河流沉积物中As浓度(55～189mg/kg)相当［34］。
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图 4 土壤中总As浓度(a)和总F浓度(b)Figure 4. The concentrations of total As (a) and F (b) in the soil samples土壤母质是土壤中氟化物的基本来源。土壤中总F浓度如图4b所示，总F浓度在1162.70～1285.10mg/kg之间，三次采样的平均值为1237.40mg/kg。表现为丰水期、平水期浓度大于枯水期，土壤总F浓度随季节变化差异不大。与西藏土壤总F背景值(542mg/kg)相比［32］，温泉淋滤的土壤总F浓度显著高于背景值，是背景值的2.28倍。与全国土壤F背景值(453mg/kg)以及世界土壤F中位值(200mg/kg)相比［35］，温泉淋滤的土壤氟化物显著偏高(P＜0.01)。与云南省洱源县高氟温泉点附近土壤总F浓度相比(630.48～1000.27mg/kg)［36］，其浓度也处于居高水平，其来源主要受高氟温泉水的冲刷、沉降和土壤吸附。水溶性的氟化物会对周边地下水和生物体产生富集影响，从而造成氟威胁。因此，对温泉水淋洗过的土壤样品进行了可溶态氟离子测定，土壤可溶态氟离子浓度在3.47～9.37mg/kg之间，表明温泉淋洗后的土壤可溶态氟浓度占比较高。
3.4 水体中金属离子浓度
以平水期为代表，测定了地热水和温泉水样品中的元素组成。结果如表5所示，地热水常量组成主要为Na和K；而温泉水中Na离子占主导，其次为Ca。通常，F离子浓度受氟石(CaF2)溶解积(Ksp)约束，Ca离子浓度越低，溶液中F离子浓度就会越高，地热水和温泉水中Ca离子浓度在6.36～35.74mg/L之间，低浓度Ca离子为F离子富集提供有利条件。此外，地热流体F离子溶解还会受到多因素的影响，如温度、pH、配体、共存离子等，使得F离子浓度分布出现反常。温泉水检出10.20mg/L的Mg离子，而在地热水却几乎未检出，Mn离子也出现了相同的规律。考虑到Mg和Mn两元素主要存在于母质矿石中，猜测温泉水与地表母质岩石中的Mg和Mn氧化物发生了交换。其他金属如Be、Cr、Pb、Sb和Mo等组分的浓度分布几乎没有显著差异。
表 5 地热水和温泉水中金属元素浓度Table 5. The metal element concentrations in geothermal and hot spring waters样品类型 金属元素浓度(mg/L) Ca K Mg Na Fe V Be Mn Cr Pb Sb 温泉水 35.74 15.46 10.20 195.4 0.091 0.009 0.002 0.14 0.035 0.0002 0.014 地热水 6.36 36.63 ＜0.013 456.5 0.047 0.013 0.005 0.012 0.049 0.0002 0.027 样品类型 Mo Cd Ti Se Zn Cu Ni Co Ba Hg 温泉水 0.035 0.0002 0.018 ND 0.002 0.0008 0.0032 0.00013 0.16 ＜0.0004 地热水 0.070 0.0001 0.025 ND 0.003 0.0014 0.0029 0.00003 0.090 ＜0.0004 注：“ND”表示未检出。 4. 结论
本文结合野外调查和室内分析对西藏羊八井地热区的地热水和温泉水水样开展了水化学、As和F浓度调查，阐明了水体水化学类型及As和F浓度变化趋势，揭示了水体As和F的来源及富集机制，评价了水体和土壤As和F超标情况和生态风险，具体结论如下：①富Na贫Ca高pH是羊八井地热水和温泉水最主要的水化学特征，为As和F离子溶出富集提供了有利条件；②地热废水的不当处理存在As和F的暴露风险，受水汽蒸发浓缩影响，温泉水As和F风险相比地热水更高；③水体As和F来源主要为水-岩浸溶相互作用，温泉淋滤促进了地表土壤As和F的富集，导致土壤总As和总F浓度均显著高于西藏土壤背景值。
本文丰富了西藏地区水环境中的As和F来源探析及环境地球化学行为，为地热水持续合理开发利用和水土环境风险评价提供依据。需要进一步完善覆盖地热流经区堆龙曲流域地表水、地下水As和F生态风险评价，加强环境监测预警。其次，需要关注温泉洗浴中As和F暴露风险。
BRIEF REPORT
Significance: It is of great significance to study the concentration characteristics and sources of As and F to understand the environmental geochemical behavior of As and F in the geothermal system and their impact on the surrounding water and soil environment. Xizang Plateau is the region of China where high-temperature hydrothermal systems are intensively distributed, and the distribution of heat flow in the Xizang Plateau ranks first in China[1]. The Yangbajing Geothermal Power Plant is the highest and largest geothermal power plant in China, setting a record for the highest reservoir temperature in China[3]. The release of As and F can cause surface water and soil environmental pollution. Although researchers have made preliminary research on the concentration of these two typical harmful elements in geothermal fluids and their impact on the surrounding water environment[9-12], it is still necessary to conduct in-depth analysis of the annual variation trend of As and F concentrations in a geothermal system, the contribution of hydrochemical characteristics to the enrichment of As and F, and the ecological risks of the surrounding soil and water environment caused by As and F leakage. On the basis of previous studies, we investigated the As and F concentrations of geothermal water, hot spring water and soil samples in the Yangbajing geothermal field, analyzed the hydrochemical characteristics of the water bodies, identified the source and enrichment mechanism of As and F in a geothermal system, and evaluated the ecological risks of As and F in water bodies and surface soils.
Methods: Three periods of field collection of geothermal water, hot spring water and soil samples were carried out in the Yangbajing geothermal field in Xizang from 2021 to 2022. Conventional physicochemical parameters were measured on site, including pH value, conductivity (EC), total dissolved solids (TDS), salinity (SAL), and temperature. The indoor analysis used atomic fluorescence spectrometer and X-ray fluorescence spectrometer to determine the concentration of total As in water and soil, respectively. The F concentration was measured using the ion selective electrode method. The pollution of As and F in water samples and soils was evaluated by comparing with the allowable maximum emission values of harmful elements in the standard and specified soil background values, respectively.
Data and Results: (1) The main hydrochemical characteristics. The conventional physicochemical parameters of water quality are shown in Table 3. The pH value of geothermal fluids is between 7.87 and 9.42. Geothermal fluids have a complex matrix, with conductivity, TDS, and salinity ranging from 1670−1882μS/cm, 1126−1340mg/L, and 914−983mg/L, respectively. The changes in water quality physicochemical parameters generally show a trend of dry season>normal season>high season. The analysis of the concentration composition of eight major anions and cations shows that cation Na+ dominates in water, anions HCO− 3 and SO4 2− dominate in hot spring water, and anions Cl− and HCO− 3 dominate in geothermal water. As the result of the Fig.E.1(a), the hydrochemical type of geothermal water is Na-HCO3∙Cl, consistent with literature[10]. Rich Na, poor Ca, and high pH are the main hydrochemical characteristics of Yangbajing geothermal and hot spring water.
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E.1. Hydrochemical composition of water sample (a), the concentrations of total As and F in water sample (b), pictures of hot spring and geothermal water (c), and the concentrations of total As and F in the soils (d).(2) Concentration levels of As and F in geothermal fluids and ecological risk assessment. The total As and F concentrations of Yangbajing geothermal water and hot spring water are shown in Fig.E.1(b). The concentrations of total As and F in geothermal water are 3.16−6.50mg/L and 15.90−17.89mg/L, respectively, which hardly changes with the change of water period. The total As concentration in geothermal water is consistent with the total As concentration in Yangbajing geothermal water reported by Zhang et al.[10] (3.54−3.56mg/L). The F concentration of geothermal water shows a decreasing trend compared to the reported Yangbajing geothermal water (18.0−18.9mg/L) and Yangyi hydropower station geothermal water (19.2mg/L)[24]. As shown in Fig.E.1(c), influenced by the evaporation of water, the total concentration of As (4.18−6.50mg/L) and F (14.56−17.89mg/L) in hot spring water are higher than those in geothermal water. The concentration of As and F shows a trend of dry season>normal season>flood season. Notably, the total concentration of As and F in waters are significantly (P＜0.01) higher than the maximum allowable emission concentrations for harmful components (0.5mg/L for As, and 10mg/L for F) in the Geothermal Resources Assessment Method (DZ40—85). Improper treatment of geothermal wastewater may pose exposure risks to As and F in the surrounding environment.
(3) Concentration levels of As and F in soils and ecological risk assessment. Fig.E.1(d) shows the total As and F concentrations of soil samples leached from the wastewater at the hot spring discharge outlet. There is no significant difference in the total As concentration between the dry season and the normal season. The total As concentration ranges from 97.6 to 99.08mg/kg, which is significantly higher (P<0.01) than the concentration during the flood season (79.5mg/kg). The total F concentration ranges from 1162.7 to 1285.1mg/kg, showing no significant variation with the seasons. Compared with the background values of total As (18.7mg/kg) and F (542mg/kg) in Xizang soil[32], the total As and F concentrations in the geothermal area are significantly higher than the background values (P<0.01), which are 4.25−5.31 and 2.28 times of the background values respectively. The results show that the soil in the geothermal area is at risk of high As and F pollution.
(4) The main sources of As and F in geothermal water and surface soil are water-rock leaching interaction, and the unique hydrochemical characteristics provide favourable conditions for the leaching of As and F. Rock leaching in the geothermal reservoir is the main resource of As and F enrichment in geothermal water. The geothermal fluids in Yangbajing have a pH value of 7.87−9.42. In addition, some hydroxyl minerals such as muscovite and biotite often undergo ion exchange with F. If the groundwater is alkaline, the exchange is more likely to occur. OH- can replace F in fluorinated minerals, increasing the concentration of F in geothermal water. It is known that the concentration of F in a water system is restricted by the solubility of fluorite. The alkaline environment caused by the hydrolysis of minerals due to water-rock interaction has a significant impact on the dissolution of fluoride, and the alkaline environment with high concentrations of Na+ and low concentrations of Ca2+ is an important reason for the formation of high fluoride. Additionally, low sulfide concentrations (as low as 16.79mg/L) further promote high arsenic geothermal water. It is worth noting that the geothermal water in Xizang has high sodium (up to 445.5mg/L), low sulfur (16.79−26.3mg/L), low calcium (3.31−6.66mg/L), and weakly alkaline (8.95−9.15), providing better convenient conditions for the dissolution of high As and F in Yangbajing geothermal fluids.
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