Heavy Metal Transport in Dammed Rivers: Damming Effects and Remediation Strategies—A Review

Abstract

Rivers, vital for life and civilizations, face severe threats from human activities such as hydropower development, with heavy metal pollution emerging as a critical concern due to altered biogeochemical cycles. Understanding how river damming affects heavy metal transport processes and developing targeted remediation strategies are essential for safeguarding the health of river-reservoir ecosystems and enabling the sustainable utilization of hydropower resources. Therefore, this review first summarizes the global hydropower development, details how damming disrupts hydrology, environments, and ecosystems, and analyzes heavy metal distribution and transport in reservoir water, suspended sediments, and riverbed sediments. It reveals that river damming promotes heavy metal adsorption onto suspended particles, deposition in riverbed sediments, and re-release during reservoir regulation, and anthropogenic activities are a primary driver of significant pollution in key reservoirs worldwide. Additionally, we further evaluate in situ (e.g., stabilizing agents, sediment capping, and phytoremediation) and ex situ (e.g., dredging, chemical washing, electrochemical separation, and ultrasonic extraction) remediation techniques, highlighting the challenges of phytoremediation in deep, stratified reservoir environments. Moreover, solidification/stabilization emerges as a promising in situ strategy, emphasizing the need for specific approaches to balance pollution control with hydropower functionality in dammed river systems.

1. Introduction

Rivers, a crucial component in the origin of life on Earth, serve as important channels in global material cycles [1]. They have nurtured and fostered the development of brilliant human civilizations and hold significant social, political, economic, cultural, and ecological value [2]. However, over the past century, rapid population growth and industrial development worldwide have posed serious threats to the health and safety of rivers [3,4]. Many large river basins are facing challenges such as water quality degradation, ecosystem destruction, and the shrinking and depletion of riverbeds [1,2]. Among these issues, heavy metal pollution in rivers has been exacerbated by unsustainable practices in mineral extraction, metal smelting, manufacturing processes, and improper agricultural activities, as well as waste discharge from transportation, industrial production, and daily life [5]. Consequently, research into the distribution pattern and transport process of heavy metals in large river basins has become a top priority in river ecosystem management.

Since the 21st century, global warming has caused significant harm to human society and economic development [2]. In an effort to control greenhouse gas emissions, countries worldwide are seeking and developing more renewable energy sources to replace traditional fossil fuels [6]. Large river basins, especially their headwater regions, possess abundant hydropower resources. According to the Renewables 2024 Report, the hydropower development has contributed over 4000 TWh of electricity globally, acting as the most predominant renewable energy source [7]. However, river damming has correspondingly presented substantial impacts on the hydrological characteristics, physicochemical environment, and ecosystems of large river basins [8]. Hydropower development alters the hydraulic conditions of rivers, impeding the transport of heavy metal pollutants, thereby creating favorable conditions for pollutant accumulation in reservoir areas [9,10]. On the other hand, processes such as water level regulation, sediment dredging, and reservoir discharges from hydropower stations change the physicochemical environment within reservoirs, leading to the suspension, transport, and release of heavy metal pollutants [11,12]. This, in turn, affects the ecological environment of both the headwater regions and downstream areas of rivers [12]. In recent years, more and more countries and regions have focused on developing cascade hydropower systems to fully utilize the hydropower resources of large river basins [13]. The dense hydropower development and river damming lead to more complex effects on the material cycles of rivers, as well as cumulative impacts on the transport of heavy metals [14]. Therefore, studying the transport processes of heavy metals under the influence of river damming holds significant scientific importance for accurately assessing the fate of river pollution.

Hydropower development has led to the deposition and accumulation of large amounts of heavy metals in reservoir sediments [5,10]. Therefore, the management of heavy metal pollution in riverbed sediments is crucial for the sustainable development of river-reservoir ecosystems [15]. To avoid secondary pollution caused by physical or chemical treatments, biological techniques are often employed for the remediation of heavy metal pollution in rivers, including methods such as plant and animal absorption, microbial regulation, and biomass stabilization [16,17]. However, unlike rivers in plains, the water environment created by high dams and large reservoirs is characterized by deep water, low flow velocity, and thermal-oxygen stratification [18,19]. These unique conditions present various challenges for the biological treatment of heavy metal pollution in reservoir sediments. Additionally, hydropower projects are often located in riverine headwater regions, where poor nutrient conditions, fragile ecosystems, and extreme mountainous environments limit the applicability of biological treatment methods [13]. Therefore, it is essential to analyze treatment methods for heavy metal pollution in reservoir sediments based on the distribution and transport characteristics of heavy metals within reservoirs.

Based on the Web of Science database, this review conducted a literature retrieval with “Dam”, “Reservoir” and “Metal, or Cadmium (Cd), Lead (Pb), Mercury (Hg), Chromium (Cr), Arsenic (As), Copper (Cu), Nickel (Ni), and Zinc (Zn)” as the themes, and obtained a total of 5261 relevant papers over the past thirty years (Figure 1A). Subsequently, CiteSpace (Version 6.4.R2) was used to perform keyword co-occurrence analysis on the 300 most relevant papers in the past five years. The results showed that keywords such as “heavy metals”, “pollution”, “water”, “contamination”, “river” and “sediment” had the highest occurrence frequencies, indicating that the issues of transport, pollution, and remediation of heavy metals in dammed rivers are of great concern (Figure 1B). Thus, this review first outlines the global status of hydropower development and the impacts of river damming on hydrological characteristics, the physicochemical environment, and ecological systems. This is crucial for understanding how damming affects the distribution and transport of heavy metals. Next, the concentrations and pollution levels of heavy metals in representative reservoirs’ water bodies, suspended sediments, and riverbed sediments worldwide are summarized. The review then focuses on analyzing the impact of dam construction on heavy metal transport, covering three key aspects: (i) Migration of heavy metals from reservoir water to suspended sediment, (ii) Deposition of heavy metals from suspended sediment to riverbed sediment, and (iii) Release of heavy metals from riverbed sediment during reservoir regulation. Finally, based on the distribution and transport of heavy metals in reservoirs, potential remediation strategies for addressing heavy metal pollution in reservoir sediments are proposed. This review aims to provide valuable insights for future management of river-reservoir systems.

2. Status of Hydropower Development and Impact of River Damming

2.1. Global Hydropower Development Status

In the late 1870s, the world’s first hydropower station was constructed and put into operation in France, sparking a wave of hydropower development across Europe and North America [2]. In recent years, as hydropower development has flourished, over 70% of the world’s major river basins have been moderately or severely modified [8]. According to the International Commission on Large Dams (ICOLD), by the end of 2024, a total of 62,339 large dams (a dam with a height of 15 m or greater from lowest foundation to crest or a dam between 5 m and 15 m impounding more than 3 million m³) had been constructed globally [20]. Among these, China, United States of America, and India have built the largest numbers, with 24,207, 10,126, and 4488 large dams, respectively. These dams collectively control approximately 3.5 trillion cubic meters of water, accounting for over 38% of the world’s total available water resources [21]. The Global Status Report on Renewable Energy (REN24-GSR) estimates that the total global hydropower capacity developed thus far is over 4000 TWh, with at least 68% of hydropower resources still awaiting development [7]. Today, with the World Bank’s resumption of loans for hydropower projects and the growing demand for hydropower resources worldwide, the number of hydropower stations is expected to continue to increase, and the total capacity of hydropower resources will steadily rise [2].

It is noteworthy that Asia possesses the richest hydropower resources worldwide, with its theoretical potential accounting for 47.2% of the global total [22]. It is home to world-renowned rivers such as the Yangtze River, the Lena River, and the Ganges River. There are 42 major rivers in Asia, whose drainage areas account for 56% of the continent’s total area. In Asia, China has the richest hydropower resources with immense development potential [23]. As of 2024, China had developed 338.7 GW of hydropower, accounting for 29% of the global total, far surpassing other hydropower-rich countries such as Brazil (9%), Canada (7%), and the United States (7%), making it the world leader in hydropower development [7]. However, the degree of hydropower resource utilization in China remains relatively low, with the developed capacity representing only 37% of the total hydropower potential, significantly lower than the over 60% development rate in many developed countries such as Europe and North America [24]. Moreover, China’s hydropower resources are unevenly distributed across regions. Approximately 80% of hydropower potential is concentrated in the southwestern region, yet only 32% of this potential has been utilized [25]. In 2020, President Xi announced Chinese “carbon peak” and “carbon neutrality” goals at the 75th United Nations General Assembly, providing new opportunities for the national energy transition and green development [26]. Currently, China has established 13 hydropower development bases, and future development efforts will focus on the southwestern region, including major rivers such as the Jinsha River, the Mekong River, the Nu River, the Yalong River, the Dadu River, the upper reaches of the Yellow River, the middle reaches of the Yarlung Zangbo River, and the main stream of the Yangtze River [23].

In addition, Central and South America ranks among the top in the world in terms of hydropower resources, with its theoretical potential making up 20.2% of the global total. The region has an extensive network of river basins, including world-famous rivers such as the Amazon River [22]. For Africa, it is also endowed with extremely abundant hydropower resources, whose theoretical potential represents 12.3% of the global total. However, it has the lowest development degree, at only around 10%, indicating enormous development potential [22]. Since 2014, the hydropower installed capacity in Central and South America and Africa has started to grow rapidly, with the total installed capacity reaching 188.4 GW and 39 GW in 2022, respectively [20,27]. Unlike the rapid hydropower development in Asia, Africa, and Central and South America, North America and Europe exhibit a gradually slowing pace. North America has relatively moderate hydropower resources, with its theoretical potential accounting for 9.2% of the global total. And Europe accounts for 6.9% of the global hydropower resources [21,22]. Recently, the growth of hydropower installed capacity in North America and Europe has been slow, and the total installed capacity stood at 200.3 GW and 309.6 GW in 2022, respectively [27].

2.2. Impact of River Damming on Hydrology, Environment, and Ecology

Since the early 20th century, the rapid development of the hydropower industry has led to varying degrees of alteration in river ecosystems worldwide, profoundly influencing their development and evolution [24]. Originally, river ecosystems were complex, open, dynamic, non-equilibrium, and nonlinear systems, characterized by continuity and integrity in their structure, function, and ecological processes [28]. However, river damming has disrupted the natural continuity and integrity of rivers, altering hydrological and hydraulic characteristics, material transport processes, and energy flow conditions [1,3]. This has resulted in the deterioration of water quality, loss of biodiversity, and damage to ecosystems.

In recent years, the ecological impacts of hydropower development on rivers have become a major focus of scientific research and have garnered attention from the World Commission on Dams and governments worldwide [4]. Based on Vannote’s River Continuum Concept (RCC), Ward and Stanford introduced the Serial Discontinuity Concept (SDC), which highlights how hydropower development disrupts river continuity, leading to stepwise changes in both biotic and abiotic processes [28,29]. Overall, the environmental impacts of river damming on water ecosystems are primarily reflected in changes to hydrological characteristics, physicochemical environment, and ecological system [8]. These changes often interact with each other, leading to regional and even global ecological transformations [30].

2.2.1. Effects of River Damming on Hydrological Characteristics

The physical barriers created by large dams alter the natural riverine shapes, leading to the reconstruction of their hydrological characteristics [31]. The construction of hydropower stations reduces the flow in downstream sections, with some tributaries even experiencing interruptions or disappearance, thereby altering the hydrological rhythms of both large and small rivers [32]. Research by Dou and Yang on the Cao’e River revealed that after damming, the annual average river flow slightly decreased, and intra-annual runoff distribution became more uniform [33]. Studies on the effect of dam operation in North Korea showed that the average monthly river flow was significantly reduced by 27.7% and 40.2% in the Imnam Dam and Hwacheon Dam, respectively [34]. Moreover, Yang et al. studied the Yenisei River in Siberia and found that the operation of upstream hydropower stations reduced downstream summer runoff by 10–50% but increased winter runoff by 45–85% [35]. Research on Chinese southwestern rivers indicated that damming reduced the multi-year average flow of rivers and led to interruptions during the dry season [36]. Guo and Yang’s study on the Sanmenxia Reservoir highlighted that post-dam, the distribution of river runoff throughout the year became more even, with increased flow during non-flood seasons [37]. Wang et al. found that reservoir operations altered the hydrological regime of rivers, particularly affecting water flow in spring and autumn, significantly reducing seasonal differences in river discharge [38]. Dang et al. reported that the cascade dams along the Mekong River, southwestern China, have shifted by ~20% of the mainstream annual volume between the dry and wet seasons in the recent decade, although the river damming has a minimal impact on the annual flow volume [39]. Therefore, hydropower development and river damming impact the seasonal variation of river flow and hydrological cycles. During the wet season, reservoir impoundment reduces the pulse effects of floods, while reservoir water releases during the dry season significantly increase river flow.

Hydropower development not only affects river flow characteristics but also alters sediment transport processes. Since the 1990s, with the large-scale construction of high dams and reservoirs, over 50% of global river sediments have been intercepted and deposited, becoming a major factor in the reduction in sediment transport in many large rivers [40,41]. River damming weakens hydrodynamics, reducing the sediment-carrying capacity of water and causing suspended sediments to settle in the reservoir areas [42]. Chen et al. pointed out that hydropower development intercepts sediment transport, leading to significant sediment deposition within high dams and reservoirs, and drastically reducing sediment loads downstream, thus altering the natural material transport processes of large rivers [43]. Studies on the Longyangxia and Liujiaxia reservoirs on the Yellow River found that damming caused a significant reduction in the annual sediment load and altered the intra-annual sediment transport pattern, with a reduction of more than 20% in sediment concentration during the wet season [44]. Moreover, Nguyen et al. found that because of the cascade damming regulation in the Mekong River, the sediment loads were decreased by 68.5% at downstream stations [45]. Therefore, for high-dam reservoirs on river channels, when high-velocity upstream waters enter the low-velocity reservoir area, the water’s sediment-carrying capacity drastically decreases, causing coarse-grained sediments to settle within the reservoir area [46].

2.2.2. Effects of River Damming on Physicochemical Environment

Although hydropower development generates a substantial amount of clean energy, its impact on the physicochemical environment cannot be overlooked. The hydrological changes are a key driving force affecting the river’s physicochemical conditions, such as water temperature, dissolved gases, nutrients, and various pollutants [8,47].

The impact of river damming on water temperature is primarily reflected in two aspects: thermal stratification within reservoirs and the release of cold water from the bottom layers [47]. The deepening of water in reservoirs due to impoundment makes it difficult for solar heat to penetrate to the bottom, leading to the formation of thermal stratification [48]. Additionally, many dam reservoirs use bottom-layer water for hydropower generation, with outlets often located below the thermocline [49]. As a result, the released water is typically colder than the downstream water, and this delayed cooling effect can alter the habitat for aquatic organisms [47]. Similarly, hydropower development significantly affects the dissolved gases in river water, mainly through gas stratification and the release of supersaturated water [50]. The surface water in reservoirs tends to have higher dissolved gas levels due to the oxygen production by phytoplankton during photosynthesis, while deeper water layers, lacking reoxygenation mechanisms, have lower levels of dissolved oxygen [51]. Moreover, in oxygen-deficient environments, the decomposition of organic matter further depletes the oxygen content, creating stagnant, hypoxic zones [50]. However, during reservoir flood releases, intense water–gas exchange can lead to supersaturation of dissolved gases in the released water. Research by Chen et al. on the Three Gorges Hydropower Station showed that when the reservoir discharge flow reached 40,000 m³/s, the gas supersaturation level peaked [52].

Although river damming and water storage cannot directly generate nutrients and pollutants, they can accumulate in large quantities due to interception, affecting the aquatic environment [32]. Covich’s global analysis of hydropower stations showed that hydropower development increased river hydraulic retention time from 16 days to 47 days, nearly tripling it [53]. With the reservoir water remaining stagnant for long periods, its self-purification capacity is much lower than that of dynamic water bodies, making it prone to eutrophication. Von Schiller’s study of the Mediterranean River revealed that while nitrogen release and storage upstream of reservoirs are balanced, nitrogen release downstream is significantly less than storage, indicating that reservoirs act as important nitrogen sinks [54]. Friedl’s research on the Columbia River in Canada showed that river damming and water storage reduced downstream phosphorus levels, leading to a significant decline in phytoplankton biomass [32]. Kuang et al. compared the water quality before and after the impoundment of Three Gorges Reservoir and found a considerable decline in water quality in the Yangtze River due to river damming [55]. Parks’ study of the Verde River in the U.S. found that over 70% of particulate organic matter was trapped in the river reservoirs [56]. The organic matter accumulating at the bottom of reservoirs undergoes anaerobic decomposition, producing acidic substances like carbon dioxide and hydrogen sulfide, which can accelerate the corrosion of dam concrete and curtains, posing potential safety risks [57]. In addition, heavy metals and emerging organic pollutants also tend to accumulate in the process of damming and water storage [58]. In the reservoir’s slow-flowing conditions, most of heavy metals are adsorbed and deposited in the riverbed sediment [5]. Gao et al.’s study found that after the Three Gorges Reservoir was impounded, the levels of Cd, Hg, Zn, Pb, and Cu in the reservoir sediment increased significantly [9]. Similarly, Wang et al.’s research on the Manwan Reservoir indicated that intensive agricultural and industrial activities in the basin led to the accumulation of As, Cd, Pb, and Zn in the reservoir sediment [59]. Thus, the prolonged hydraulic retention time in reservoirs creates favorable conditions for the accumulation of nutrients, heavy metals, and other pollutants.

2.2.3. Effects of River Damming on Ecological System

Rivers serve as essential habitats for the distribution, habitation, and reproduction of fish, benthic organisms, and plankton [60]. Hydropower development alters the hydrological characteristics and physicochemical environment of rivers, thereby impacting the survival and development of aquatic organisms. These changes can lead to shifts in the structure and functioning of river ecosystems.

The impact of hydropower development on river fish is mainly manifested in the displacement of spawning grounds, obstruction of migratory routes, and habitat destruction [61]. The impoundment of reservoirs can submerge fish spawning areas, while water release can erode them, forcing fish to seek new spawning locations [62]. Additionally, the physical barriers created by high dams disrupt the natural continuity of rivers, hindering both feeding and reproductive migrations of fish [63]. For example, the Edwards Dam in the U.S., significantly reduced fish biomass by obstructing the migration and spawning of species like salmon and bass, leading to ecosystem degradation. Furthermore, structures associated with high dams and reservoirs, such as spillways and turbines, can also adversely affect fish habitats. These effects include high-pressure, high-velocity water impacts, cold water releases, and the discharge of gas-supersaturated water [47,50]. For instance, in the Columbia River in the U.S., gas-supersaturated water caused by dam spillways during the flood season leads to the death of large numbers of juvenile fish each year [50].

Benthic organisms, which often reside on riverbed rocks or aquatic vegetation, play a crucial role in nutrient cycling and energy flow within aquatic systems [64]. However, hydropower development significantly impacts their survival by altering river flow velocity, water depth, temperature, and nutrient conditions. After reservoir impoundment, the reduced water flow velocity affects the lateral migration of benthic invertebrates, while increased water depth changes their vertical distribution [65,66]. Additionally, reservoir impoundment leads to thermal stratification, with cold water releases lowering the overall river temperature, which can impact the types and abundance of benthic organisms. Furthermore, the construction of hydropower stations can also affect benthic habitats. Chen et al. found that after damming and water storage, the differences in substrate conditions diminished, leading to a decrease in the abundance, density, and diversity of benthic invertebrates and simplifying community structures [67]. Hydropower development also has significant effects on benthic plant communities. Research by Maavara et al. showed that reservoir impoundment can submerge large amounts of emergent vegetation, while changes in river hydrological characteristics, nutrient structure, and light conditions have substantial impacts on the growth of submerged aquatic plants [1].

Plankton, which includes both zooplankton and phytoplankton, thrives suspended in the water column and is easily transported passively by wind and water currents [68]. Phytoplankton, in particular, are sensitive to changes in the aquatic environment, with algal blooms serving as a key indicator of water quality degradation and ecological damage in reservoirs [69]. Research by Zhou et al. on the Xiangxi River Reservoir found that the closed, stagnant conditions of reservoirs are conducive to triggering algal blooms, whereas in flowing river environments, such blooms typically occur only during the warmer, sunnier spring season [70]. Similarly, Jia et al.’s study on algal communities in the Gufu River hydropower station highlighted that changes in river hydrodynamic conditions and aquatic environments are significant factors influencing shifts in algal communities [71]. Overall, the alterations in hydrological characteristics and physicochemical environment induced by hydropower development have a substantial impact on the types, abundance, and structure of aquatic organisms.

As is well-known, dam impoundment alters river ecosystems and also results in changes in carbon cycling processes, thereby exerting a certain impact on global climate change [72]. Studies have shown that river damming and water storage promote the gradual evolution of the ecosystem’s production mode from a “river-type” heterotrophic system dominated by benthos to a “lake-type” autotrophic system dominated by phytoplankton, thereby strengthening the carbon sequestration effect of dissolved organic matter (DOM) in reservoirs [8]. However, this phenomenon mostly occurs in small and medium-sized mixed reservoirs; for large stratified reservoirs, the autotrophic system in the epilimnion and the heterotrophic system in the hypolimnion often coexist [73]. Previous studies have mainly focused on the carbon sequestration effect of DOM in the epilimnion, while greatly neglecting the carbon emission impact caused by the discharge of greenhouse gases such as CO₂ and CH₄, produced by heterotrophic respiration of DOM in the hypolimnion [74]. Existing studies have found that the carbon emission effect of the hypolimnion in stratified reservoirs is severely underestimated, and its carbon emissions can account for 90% of the total annual carbon emissions of the reservoir [75]. Therefore, future studies need to further investigate the DOM cycle in the hypolimnion of stratified reservoirs and in-depth explore its objective impacts on global carbon emissions and climate change.

3. Concentration Distribution of Heavy Metals in Water and Sediments

Heavy metals are defined as metals with a density greater than 5 g/cm³ or an atomic number greater than 20 [5]. Common heavy metal pollutants in the environment include Cd, Pb, Hg, Cr, As, Cu, Ni, and Zn. These pollutants originate from both natural sources, like rock weathering, forest fires, and volcanic eruptions, and anthropogenic activities, including the discharge of wastewater, sewage, and solid waste [76]. Among these, key point sources of heavy metals include mining, smelting, electroplating, manufacturing, and processing industries. Moreover, major non-point sources encompass agricultural and forestry activities, atmospheric deposition, and transportation [13,77,78].

In recent years, due to population growth and industrial development, heavy metal pollution in water environments has intensified. Heavy metals with significant biological toxicity, such as Hg, Cd, and Pb, can cause severe degradation of aquatic environments when discharged into rivers [79]. Unlike organic pollutants, which can be biologically degraded, heavy metals tend to persist in the environment for long periods and accumulate through the food chain, ultimately entering the human body and posing health risks [80,81]. Due to their non-degradability, toxicity, and bioaccumulation potential, heavy metals are classified as toxic and harmful substances by countries and regions worldwide [13]. Short-term exposure to heavy metal pollution can lead to acute poisoning symptoms, including headaches, memory loss, mental confusion, gastrointestinal discomfort, visual disturbances, and allergic reactions. Long-term exposure, however, can have more severe consequences, such as increased cancer risk, disruption of gene expression, and developmental disorders [80].

3.1. Heavy Metals in Reservoir Water

The issue of heavy metal pollution in the water bodies of large reservoirs directly affects the safety of drinking water and agricultural irrigation within the basin [82]. Research indicates that anthropogenic sources, such as industrial and domestic wastewater discharge, are significant contributors to the increased concentration of heavy metals in reservoir waters (

Table 1. Concentrations of heavy metals in worldwide reservoir waters.

Large Reservoir	Concentrations of Heavy Metals (µg·L−1)
	Hg	Cd	Pb	Zn	As	Cu	Ni	Cr
Reservoir group in hilly area of South China [82]		0.03 *	0.45	51.9 *	0.95	1.68	0.96 *	1.7
Reservoir group in the Jinsha River, China [83]		0.01–1.95	1.88–69.26		1.08–3.54	4.54–35.56	0.53–65.97	3.56–38.42
Reservoir group in the Mekong River, China [84]		0.01–0.03	0.40	10.6	5.0–6.6	0.35–0.61	0.03–0.23	0.14–0.36
Three Gorges Reservoir in the Yangtze River, China [9]	0.03	1.02	11.2 *	10.43	2.33 *	8.94
Keshan Reservoir in the Yellow River, China [85]	0.05 *	0.5	2.2	0.06	1.7	0.02		5.8
Yuqing Reservoir in the Yellow River, China [85]	0.06 *	0.4	2.5	0.02	1.4	0.02		5.6
Manwan Reservoir in the Mekong River, China [86]	0.2 *	3.9	32	706	21	80
Xiaolangdi Reservoir in the Yellow River, China [87]			1.08	16.1	3.51	2.93	4.34	5.61
Danjiangkou Reservoir, China [88]		1.17	10.6 *	2.02	11.1 *	13.3	1.73	6.3
Alzate Reservoir, Mexico [89]	104 *		61 *	68		70	34	79 *
Iron Gate Reservoir, Serbia [90]		1.1	8.3	85.4		65.1
Kralkizi Reservoir, Turkey [91]		0.036	2.56	5.02	2.39	2.83	15.8	22.1
Dicle Reservoir, Turkey [91]		0.030	1.84	4.12	1.61	2.12	15.9	18.6
Batman Reservoir, Turkey [91]		0.044	1.56	4.09	0.71		16.0	16.5
Atatürk Reservoir, Turkey [92]				64		25.0	15.4

Note: * indicates significant anthropogenic pollution of heavy metals.

).

Wang et al. found that due to varying degrees of human disturbance, heavy metal concentrations exhibit high spatial heterogeneity among reservoir clusters in the hilly regions of southern China [82]. In Hunan Province, wastewater discharge from electroplating, mining, and dyeing industries has led to severe exceedance of Zn levels in reservoirs [82]. In the Three Gorges Reservoir, industrial emissions and gasoline combustion have resulted in significantly higher Pb contents in the reservoir water compared to Chinese surface water environmental quality standards. Additionally, frequent vanadium-titanium mining in Panzhihua City has exacerbated Pb and As pollution in the Three Gorges Reservoir area [9]. In the lower reaches of the Yellow River, Hg pollution is notably severe, with coal-fired power plant emissions contributing to the formation of toxic methylmercury compounds in the water environment, impacting aquatic health [85]. In Guizhou, an organic chemical plant has directly discharged Hg pollution into the Baihua Reservoir, resulting in significant organic Hg contamination in the reservoir [93]. The Danjiangkou Reservoir has also experienced severe As and Pb pollution due to industrial and agricultural wastewater discharges [88].

In addition, there are many large reservoirs where heavy metal concentrations are relatively low, reflecting the geological background of riverine heavy metal levels. For example, the Atatürk, Kralkızı, Dicle, and Batman reservoirs in Turkey, as well as the Iron Gate Reservoir in Serbia, have never exceeded the heavy metal concentration thresholds set by the World Health Organization (WHO), European Union (EU), Environmental Protection Agency (EPA), and China’s Ministry of Ecology and Environment [90,91,92]. Moreover, in the reservoir clusters of the hilly regions of southern China, the unique red soil composition has led to significantly higher aluminum concentrations in the reservoir water. This indicates that the geological background of the reservoir area is also an important factor influencing heavy metal levels in the water [82].

3.2. Heavy Metals in Reservoir Suspended and Riverbed Sediments

Generally, the concentration of heavy metals in reservoir suspended and riverbed sediments is notably higher than in water bodies, as in many cases over 99% of heavy metals are adsorbed onto particulate matter and eventually deposited in sediments [5,9]. Thus, sediments not only act as a major “sink” for heavy metals but can also become a pollution “source” that affects the water quality when hydrodynamic processes and biological disturbances occur [10]. Therefore, in recent decades, many researchers have focused their studies on the concentrations of heavy metals in suspended and riverbed sediments of large reservoirs (

Table 2. Concentrations of heavy metals in worldwide reservoir sediments.

Large Reservoir	Concentrations of Heavy Metals (mg·kg−1)
	Hg	Cd	Pb	Zn	As	Cu	Ni	Cr
Riverbank zone, Three Gorges Reservoir, China [94]		0.99 *	57.1 *	161 *		69.3 *
Riverbed zone, Three Gorges Reservoir, China [94]		0.88 *	51.0 *	174 *		54.2 *
Three Gorges Reservoir, China [95]	0.17 *	0.90 *	44.0 *	130.3 *	14.1	56.4 *	45.7	84.9
Three Gorges Reservoir, China [96]	0.13 *	0.87 *	48.1 *	161.3 *	14.3	58.4 *	51.1	105.4
Reservoir group in the Jinsha River, China [83]		0.3–4.6 *	4.9–164 *		0.2–9.1	9.2–280 *	4.0–215	12.8–135
Manwan Reservoir in the Mekong River, China [59]		1.41 *	47.1 *	156.7 *	40.6 *	38.9		54.70
Yuqiao Reservoir in the Huaihe River, China [97]		0.30 *	24.0	83.9		31.6	29.8	59.40
Xiaolangdi Reservoir in the Yellow River, China [98]			31.6	124		28.2	32.7	65.3
Rybnik Reservoir, Poland [99]		25.8 *	118.6 *	1583 *		451 *	71.1 *	129.8
Iron Gate Reservoir, Serbia [90]		7.85	35.4	114.3		67.9
Iron Gate Reservoir, Serbia [100]	0.23	3.0 *	43.6	307.8 *	9.2 *	57.6 *	74.5 *	93.3 *
Iron Gate Reservoir, Serbia [100]	0.19	2.1	28.0	197.5	3.2	31.6	59.2	71.1
Paiva Castro Reservoir, Brazil [101]		0.3	13.5	12.9		3.9	1.4
Rio Grande Reservoir, Brazil [102]		9.7 *	765.9 *	128.6		1644 *	74.4	56.7
Atibainha Reservoir, Brazil [103]			21.89	52.26		23.9	10.92	33.40
Igarata Reservoir, Brazil [103]			23.83	58.78		16.3	12.19	43.53
Itupararanga Reservoir, Brazil [103]			24.53	39.80		18.6	7.15	30.88
Barra Bonita Reservoir, Brazil [103]			18.53	78.28		51.1	40.6 *	34.12
Broa Reservoir, Brazil [103]			15.06	42.27		32.2	12.67	25.44
Salto Grande Reservoir, Brazil [103]			21.53	88.05		51.3	23.83	45.15
Rio Grande Reservoir, Brazil [103]			42.41	138.6		66.0 *	14.65	57.04
Aguamilpa Reservoir, Mexico [104]	0.04	0.27 *	13.6	51.8		60.8 *	189 *	18.3
Vaussaire Reservoir, France [10]		0.4	19.3	159.4	9.4	27.2	58.3	157.6
Ain Reservoir, France [105]		<0.2	9.1	26.9	10.6	4.1	9.4	21.3
Bienne Reservoir, France [105]		<0.2	11.5	52.4	4.6	24.0	10.8	14.9
Ain Reservoir, France [105]		0.2	32.0	78.2	4.9	32.2	16.9	32.8
Bienne Reservoir, France [105]		0.3	23.9	85.7	5.9	45.3	6.5	12.8
Villerest Reservoir, France [106]		1.7 *	87.8 *	217.0 *	44.7 *	52.3 *	39.4	116.0
Wettingen Reservoir, Switzerland [107]		0.4 *	34.4 *	121.5 *		37.5 *	36.4	25.3
Klingnau Reservoir, Switzerland [107]		0.3	49.0	126.0		42.0	35.0	60.0
Wohlen Reservoir, Switzerland [107]		0.3	30.9	141.0		53.0	28.5	45.0
Verbois Reservoir, Switzerland [107]		0.2	19.5	57.5		22.6	33.8
Malter Reservoir, Germany [108]		22 *	420 *	1300 *		200 *		200 *
Kapulukaya Reservoir, Turkey [109]		0.9 *	21.4	45.8 *	19.4 *	19.2 *	65.8 *	327 *
Thisavros Reservoir, Greece [110]	0.03	0.13 *	38.41	59.3 *	0.64	35.00	16.7	18.0
Platanovrisi Reservoir, Greece [110]	0.04	0.09	8.22	105.5	0.25	18.04	19.40	31.95
Kafrain Reservoir, Jordan [111]		10.7 *	132.4 *	64.1 *		22.9	84.7	82.6
Mujib Reservoir, Jordan [112]		6.3 *	55.5 *	278.4 *		55.5 *	37.9	114
Kapshagay Reservoir, Kazakhstan [113]		0.46	5.18	33.6		0.23

Note: * indicates significant anthropogenic pollution of heavy metals.

).

Human activities have led to varying degrees of heavy metal pollution in suspended and bottom sediments of reservoirs globally [106,114]. For example, regarding the Three Gorges Reservoir in the Yangtze River, China, significant enrichment of heavy metals such as Cd, Hg, Zn, Pb, and Cu was observed in the sediments after the reservoir’s filling [94,95,96,115,116]. Similarly, the Manwan Reservoir in the Lancang River, China showed increased concentrations of As, Cd, Pb, and Zn due to industrial and agricultural pollution [59]. In Brazil, the Paiva Castro Reservoir exhibited a Cu concentration in the riverbed sediments that was four times its background level, a result of Bordeaux mixture used to control algal blooms, though the bioavailability of Cu is low as it binds primarily with clay and manganese (hydr)oxides [14,101,117]. The Grande Reservoir in Rio de Janeiro also experienced significant Cu, Cd, and Pb pollution due to urban and industrial waste as well as copper-based algaecides [102,103]. In Poland, the Rybnik Reservoir’s bottom sediments revealed elevated levels of Zn, Cu, Ni, Pb, and Cd due to industrial discharge and atmospheric deposition [99]. These examples underscore the significant impact of human activities on heavy metal concentrations in reservoir sediments worldwide.

However, some large reservoirs exhibit relatively low concentrations of heavy metals in their sediments, such as the Iron Gate Reservoir in Serbia and the Vaussaire Reservoir in France, primarily due to limited human pollution in these areas [10,90,105]. Compared to the heavy metal concentrations in sediments from the industrial zones of the Danube River, the Iron Gate Reservoir shows significantly lower levels [90]. Similarly, the heavy metals in the Vaussaire Reservoir primarily originate from natural materials associated with the parent soil material, with minimal human impact. Additionally, the substantial input of fresh organic matter has diluted the heavy metal concentrations in the Vaussaire Reservoir’s sediments [10]. Therefore, in large reservoirs, heavy metal concentrations in sediments typically remain at background levels of the river, with increases in concentration primarily resulting from human activities and pollution inputs.

4. Influence of River Damming on the Transport of Heavy Metals

4.1. Migration of Heavy Metals from Reservoir Water to Suspended Sediment

For dissolved heavy metals, the low-flow environment created by high dams can promote the adsorption and binding of metal ions to suspended particles, thereby reducing water heavy metal contents (Figure 2) [9,96,116]. Regarding the Three Gorges Reservoir in the Yangtze River, Gao et al. found that the concentration of dissolved heavy metals decreased spatially towards the dam [9]. Similarly, Varol et al. observed in the Kralkızı, Dicle, and Batman reservoirs in Turkey that the lowest concentrations of dissolved heavy metals were near the dam walls, while the highest were at the reservoir entrance [91]. For the Manwan Reservoir in the Lancang River, Zhao et al. noted significant spatial variations in the concentrations of iron, manganese, Ni, Pb, and Zn, with the highest levels in the central reservoir and the lowest near the dam [86]. Furthermore, the heavy metal pollution characteristics of riverine water also change with the construction of hydropower stations. Before the construction of the Manwan Reservoir, the primary pollutants in the Lancang River were Hg, Pb, and Cu, but after construction, Hg, Cd, and Pb became the main pollutants [86]. Additionally, the seasonal operation of high dams affects the concentrations of dissolved heavy metals. Reports indicate that during the reservoir filling period, the concentrations of heavy metals such as magnesium, titanium, vanadium, cobalt, As, and Cd are significantly higher compared to the flood discharge period [118]. This is attributed to increased heavy metal pollution input into the reservoir water during the wet season due to intense precipitation and soil weathering [119].

4.2. Deposition of Heavy Metals from Suspended Sediment to Riverbed Sediment

Under the influence of high dams, the reduced flow velocity in reservoirs decreases the river’s sediment transport capacity, thereby increasing sediment retention time and promoting the settlement of sediment particles [32,94]. Coarse sediment particles typically settle at the reservoir tail, forming sedimentary deltas, while finer particles are carried toward the reservoir center, and the smallest particles and colloids eventually reach the reservoir head through density flows [94]. Thus, the distribution of suspended sediments in the reservoir separates according to particle size and density, with the smallest particles at the reservoir head, which facilitates the adsorption and settling of heavy metal pollutants (Figure 2) [96,120].

Regarding the Three Gorges Reservoir in the Yangtze River, China, Wang et al. studied the distribution of heavy metals in both riverbank and riverbed sediments and found that Cr and Ni concentrations increased towards the dam, with the highest concentrations at the reservoir head [121]. Bing et al. observed that Cd concentrations in riverbank sediments increased towards the reservoir head, while there were no significant differences in heavy metal concentrations in riverbed sediments between the head, middle, and tail of the reservoir, which is attributed to differences in environmental conditions between riverbank and riverbed sediments [94]. In the reservoir’s riverbank areas, various hydraulic adjustments and human activities directly increase water erosion intensity, altering heavy metal distribution, whereas the water dynamics in the riverbed’s deep water environment remain relatively stable, avoiding external hydraulic disturbances and internal pollution releases [94]. In the Manwan Reservoir, Wang and Zhao found that heavy metal concentrations at the reservoir head were significantly higher than those in the middle and tail, indicating a significant impact of high dams on sediment heavy metal distribution [59,120]. García-Ordiales et al. studied the Castilseras Reservoir in Spain and found that due to selective accumulation of fine particles, heavy metal concentrations in the sediment near the head of the dam were notably higher [122]. Conversely, the interception of sediment transport by high dams results in relatively lower heavy metal concentrations downstream. Papastergios et al. investigated the Thisavros and Platanovrisi reservoirs in the Nestos River in Greece and found that Cd and Zn pollution predominantly settled within the reservoir, significantly alleviating ecological damage caused by heavy metal pollution in downstream areas [110]. Kang et al. found that hydropower reservoirs enhanced the accumulation of heavy metals towards surface sediments and aggravated ecological risks in Jiulong River Basin, China [123].

4.3. Release of Heavy Metals from Riverbed Sediment During Reservoir Regulation

Although hydropower development can lead to the accumulation and enrichment of heavy metal pollutants within a reservoir, frequent reservoir management activities, such as out-of-season water level adjustments, sediment dredging, and reservoir discharge, can also cause the mobilization and release of these pollutants [11]. Changes in the reservoir’s physicochemical conditions, such as pH, suspended particulate matter concentration, and redox conditions, significantly affect the migration of heavy metals within the reservoir [12].

The seasonal water level adjustments in high dam reservoirs cause alternating wet and dry conditions along the riverbanks, which can lead to repeated cycles of heavy metal release and deposition, altering the distribution of pollutants [124]. Shotbolt et al. observed that when shear forces from water flow exceed the cohesive forces between sediment particles, heavy metals deposited in the reservoir can become suspended and migrate [125]. Vukovic et al. reported that water level adjustments in the Iron Gate Reservoir on the Danube River result in changes in hydrological characteristics and sediment transport, impacting the fate of heavy metals [90]. During low-flow periods, heavy metals tend to accumulate with sediments, while during high-flow periods, hydraulic erosion can cause the re-suspension of sediment-bound heavy metals [90]. In the Three Gorges Reservoir in the Yangtze River, the repeated fluctuations in water levels lead to the movement and migration of heavy metals along the riverbanks, posing potential threats to the downstream aquatic environment [126].

Reservoir sediment flushing and dredging are critical measures for maintaining the storage capacity of large reservoirs (Figure 2) [11,12]. During these operations, the reservoir may partially or fully open the outlet to rapidly lower the water level to the minimum operating level and flush sediments through the bottom valves, discharging them downstream [127]. However, these reservoir management processes affect water parameters such as pH, suspended particle concentration, redox conditions, temperature, and salinity, which in turn influence the migration and transformation of heavy metals [128]. Research indicates that during sediment dredging, the oxidation of sulfides or degradation of organic matter can release heavy metals from sediments, posing potential ecological risks [11]. Additionally, Eggleton and Atkinson have noted that water parameters like particle concentration, pH, and redox conditions affect the migration of heavy metals between dissolved and particulate phases through complexation or competitive effects, thereby altering their bioavailability [129,130]. Frémion’s laboratory experiments on sediment-heavy metal desorption equilibrium show that suspended particle concentration is a key factor affecting the migration of heavy metal pollution [12]. Bing’s research found that sediment particle size and the content of iron and manganese oxides are crucial factors controlling heavy metal migration [94]. Chen et al. investigated the influence of artificial flood event on the transport of particulate heavy metals from Xiaolangdi reservoir and found that the binding of Cr, Ni and Cu with iron-manganese (hydr)oxides was enhanced during the sediment delivery, and these heavy metals were greatly scavenged by reservoir-sourced fine particles [131].

5. Remediation Strategies of Heavy Metals Polluted Reservoir Sediments

In river-reservoir systems, sediment heavy metal pollution control technologies are primarily categorized into two types, in situ and ex situ remediation techniques [5,15].

5.1. In Situ Treatment Techniques

In situ remediation technologies focus on stabilizing and immobilizing heavy metals in sediments through physical, chemical, and biological methods. Specific approaches include sediment washing and separation, agent stabilization, sediment capping, phytoremediation, and microbial remediation [132]. These technologies are typically applied directly at the contamination site, making them relatively cost-effective [133]. Stabilizing agents often have high cation exchange capacities to reduce the chemical mobility and bioavailability of heavy metals. Sediment capping involves covering contaminated sediments with clean sediments to minimize direct interactions between the water and the contaminated sediments, thus isolating heavy metals from aquatic environments both physically and chemically. Bioremediation, for example, phytoremediation uses biological organisms to absorb, accumulate, or transform heavy metals, thereby removing contaminants from the sediment environment (Figure 3) [5].

5.1.1. Stabilizing Agents

Stabilizing agents typically possess high cation exchange capacity, which reduces the chemical mobility and bioavailability of heavy metals in sediments through adsorption and precipitation processes [134]. Additionally, these agents need to have high adsorption capability, low water solubility, and high redox stability [5]. For in situ stabilization of heavy metals, inexpensive materials such as apatite, zeolites, and steel shot are commonly used [135]. Apatite is an ideal sediment treatment material. During the sediment treatment process, metals first undergo ion exchange with calcium and enter the apatite lattice, which then stimulates the dissolution of apatite and releases phosphates [134]. Since metal ions and phosphates have low solubility, their reaction forms metal-phosphate crystals, effectively stabilizing heavy metals in the sediments [136]. He et al. found that nano-hydroxyapatite can effectively stabilize Pb and Cd in contaminated agricultural soils [137]. Gray et al. used lime and red mud to stabilize heavy metal pollution in soils and sediments around a lead-zinc smelting plant in the UK, discovering that an addition of 5% red mud can efficiently stabilize heavy metals such as Zn, Cd, Pb, Ni, and Cu [138]. Yuan et al. applied nano-chlorapatite modified biochars to immobilize Cd pollution in the Jinsha River reservoir sediments, and found that the addition of 5% modified biochars facilitated the precipitation of Cd₅(PO₄)₃Cl, Cd₃(PO₄)₂, and Cd₄P₂O₉, reduced Cd mobility and bioavailability, and even decreased Cd resistant gene abundances in prokaryotic communities [139]. To sum up, stabilizing agents are suitable for remediating moderately and lightly heavy metal-polluted sediments, and they generally exhibit moderate economy and high effectiveness (Figure 3).

5.1.2. Sediment Capping

Sediment capping technology involves covering contaminated sediments with clean sand or gravel to prevent direct contact between water and sediment, thereby reducing the dissolution and release of heavy metals (Figure 3) [18]. By providing physical isolation, the initially contaminated sediments become cleaner. Studies indicate that the optimal thickness for sediment capping is 50 cm, and this method can reduce the heavy metal content in the water by up to 80% [18]. However, this technique primarily reduces the release of heavy metals from sediments and has limited effectiveness in immobilizing them. To enhance the stabilization capability of sediment capping, many environmental scientists have explored adding stabilizing agents, such as apatite or zeolites, to the sediment [15]. These active substances can stabilize dissolved heavy metals and improve the effectiveness of sediment capping. Jacobs et al. found that adding natural zeolite to sediment significantly enhances the stabilization of heavy metals and organic pollutants [140]. Thus, as recommended by the U.S. Environmental Protection Agency (EPA), it is more recommended to use mixed stabilizing agents and reduce the capping thickness to 10~30 cm, which can further reduce the cost.

5.1.3. Phytoremediation

Phytoremediation is a technique that uses plants to extract, absorb, or stabilize heavy metals from contaminated environments [16]. This method has shown effective results in managing metals such as Zn, Cd, and As in sediments [141]. Phytoremediation involves two main components, the plants themselves and the rhizosphere microorganisms that convert toxic compounds into less harmful forms [142]. Aquatic plants absorb and accumulate various heavy metals through processes involving plant chelators and metal-binding proteins [143]. However, studies have shown that the overall metal uptake by aquatic plants tends to be relatively low [15]. To enhance phytoremediation efficiency, many researchers have employed plant growth-promoting rhizobacteria in conjunction with aquatic plants. These bacteria improve plant growth, enhance the surrounding rhizosphere environment, and activate insoluble metals [16]. For example, Plociniczak et al. used the Rhizospheric bacterial strain Brevibacterium casei MH8a, which secretes deaminase and plant hormones, to promote the absorption of Cd and As by white mustard [144]. Similarly, Wang et al. employed the endophytic bacterium Serratia marcescens PRE01, which secretes deaminase, iron carriers, and plant hormones, to enhance Cd and Cr uptake by Pteris vittata [145]. Nevertheless, the efficiency of phytoremediation is the lowest among all in situ remediation technologies, which is mainly attributed to its long remediation cycle and limited heavy metal uptake capacity [146]. Thus, phytoremediation is suitable for low-level heavy metal polluted sediments or can serve as a supplementary technology to other remediation methods, to give full play to its advantage of low cost (Figure 3).

5.2. Ex Situ Treatment Techniques

Ex situ remediation technologies focus on extracting and separating heavy metals from dredged sediments using physical, chemical, and biological methods [147]. These technologies are typically applied off-site, effectively removing a large proportion of chemically mobile heavy metals, but they also cause structural damage to the sediments. Moreover, the high cost of these technologies limits their application in large-scale polluted areas [148]. Regarding the ex situ remediation framework, physical-chemical treatment methods include washing and leaching, as well as electrochemical separation. Washing and leaching involve treating the dredged sediments with acidic solutions, chelating agents, or surfactants. Electrochemical separation is another physical-chemical technique that drives metal cations in the sediments to migrate toward the cathode and anions toward the anode, with the accumulated ions subsequently extracted via precipitation, electroplating, or ion-exchange resins. Additionally, chemical stabilization methods, such as encapsulation, vitrification, and chemical oxidation-reduction, are used to stabilize heavy metals in the sediments. Biological extraction methods involve using animals, plants, or microorganisms to extract and transfer heavy metals. Thermal treatment technologies include incineration, pyrolysis, desorption/vapor extraction, and thermal aeration. Ultrasound-assisted extraction uses ultrasonic cavitation to extract heavy metals from sediments [5].

5.2.1. Sediment Dredging

Sediment dredging technology effectively removes heavy metal contamination from river environments. However, the dredged sediment typically requires further treatment to achieve harmlessness, reduction, stabilization, and resource recovery. Harmlessness involves eliminating pathogens, parasitic eggs, and viral particles present in the sediment. Reduction refers to decreasing the water content in the sediment and achieving mud-water separation as much as possible. Stabilization aims to reduce the chemical mobility of heavy metals, while resource recovery seeks to maximize the utilization value of the sediment [148]. The coupled process of cutter suction dredging, mechanical dewatering, and solidification proposed by Peng et al. allows for environmentally friendly sediment dredging. The solidified heavy metal-contaminated sediment can be repurposed as construction fill, thus achieving resource recovery of the sediment [149]. Typically, sediment dredging is used to treat highly contaminated sediments, and since it requires combination with other remediation technologies, it often exhibits moderate to low economy and moderate effectiveness (Figure 3).

5.2.2. Chemical Washing

Chemical washing is a straightforward and practical ex situ remediation technique that involves transferring heavy metals from dredged sediment into a liquid phase by adding washing solutions. To enhance the washing effectiveness, inorganic acids like nitric acid, organic acids like oxalic acid, chelating agents like ethylenediaminetetraacetic acid (EDTA), and surfactants like rhamnolipids are used [150]. These agents aid in dissolving, dispersing, and desorbing heavy metals from the sediment, facilitating the release of exchangeable, carbonate-bound, oxidizable, and reducible metal forms [15]. Loser et al. utilized various sulfuric acid leaching methods to extract and remove heavy metal contamination from sediment in the Weisse Elster River in Germany, finding that non-biological suspension leaching had the fastest rate, while solid bed leaching had the highest efficiency [151]. Nystroem et al. discovered that among all sediment washing agents, hydrochloric acid achieved the highest desorption efficiencies for Cu, Zn, Pb, and Cd, with rates of 48%, 80%, 96%, and 98%, respectively [152]. Therefore, it can be observed that chemical washing exhibits high removal efficiency, which is suitable for treating highly to moderately contaminated sediments. However, it is also accompanied by high treatment costs and additional wastewater treatment issues [15,152]. On the other hand, chemical washing may cause the loss of nutrient components in sediments. Therefore, a comprehensive evaluation of the practical application of this method is necessary in the future (Figure 3).

5.2.3. Electrochemical Remediation

Electrochemical remediation involves applying an electric current to create a potential gradient that facilitates the migration of charged ions in contaminated sediment. Cationic pollutants, such as Ni, Cd, and Pb, are attracted to the cathode, while anionic pollutants, such as chromates and arsenates, move toward the anode [153]. After remediation, the accumulated pollutants on the electrodes can be extracted through electroplating, co-precipitation, or ion exchange. In sediment environments, fine particles exhibit the highest conductivity and adsorb most of the metal ions, making electrochemical remediation particularly effective for fine-grained sediments [153]. Research by Matsumoto et al. demonstrated that using electrochemical technology on lake sediments could achieve removal rates of up to 98% for Cd and 86% for Zn [154]. Kirkelund et al. applied electrochemical remediation to remove heavy metal pollution from harbor sediments, achieving removal rates of 98% for Cd, 78% for Zn, and 65% for Pb within two weeks under low pH conditions and a current density of 1.0 mA·cm⁻² [155]. Therefore, the electrochemical remdediation method is mainly suitable for highly heavy metal-contaminated sediments and exhibits high efficiency. However, its applied voltage also additionally increases the treatment cost (Figure 3).

5.2.4. Ultrasonic Extraction

Ultrasonic extraction employs high-energy sound waves to induce cavitation, leading to the formation, growth, and collapse of bubbles in the liquid [15]. This process generates localized high-temperature and high-pressure hotspots that facilitate the release of heavy metal elements from sediments [156]. Ultrasonic extraction is particularly effective for separating coarse-particle heavy metal contaminants but is less efficient for removing fine particles, such as those found in clay. Compared to traditional remediation techniques, ultrasonic extraction offers rapid metal pollutant removal, significantly reducing the time required for remediation. However, similar to the electrochemical remediation, it relies on external ultrasound, thus leading to relatively high treatment costs (Figure 3).

5.3. Remediation Technologies for Reservoir Sediments

In the context of hydropower development, most heavy metal contaminants bind with sediment in reservoirs and subsequently settle in the riverbed sediment [9]. In situ remediation techniques are commonly employed for managing heavy metal pollution in reservoir sediments due to their minimal interference with river hydrology, preserving the flow, water levels, and head requirements for hydropower development, and their relatively low cost [31]. Furthermore, the vast volume of accumulated sediment in reservoirs makes ex situ remediation challenging and costly, as it involves dredging the sediment, which complicates the restoration process and has significant socio-economic impacts [46,47]. Additionally, the heavy metal pollution in sediments of large dams generally originates from upstream industrial and agricultural wastewater and waste, resulting in relatively low overall pollution levels due to dilution by river sediments [10,11]. In situ remediation techniques are often sufficient for addressing these levels of contamination. Zoumis et al. studied heavy metal pollution in Germany’s Mulde Reservoir and found that Zn and Cd in sediments could be oxidatively released due to biological disturbances or hydrological regulation. They proposed coupling sediment capping with zeolite stabilization to prevent re-oxidation of sediments, effectively managing heavy metal pollution [18]. Wu et al. conducted research on China’s Dahuofang Reservoir and highlighted the urgent need for in situ remediation to address heavy metal pollution from industrial and agricultural activities [157]. Huang et al. investigated pollution in 142 lakes and reservoirs in China and noted the increasing severity of Cr, Cd, and As contamination. They emphasized the need to enhance the stabilization of heavy metal pollution in lake and reservoir sediments, in addition to controlling pollution emissions [19].

In addition, the hydrological characteristics, environmental conditions, and management practices of large reservoirs impose further limitations on in situ heavy metal remediation. The seasonal water level fluctuations caused by hydropower development can disrupt riparian vegetation ecosystems, preventing many emergent aquatic plants from stabilizing and growing in the reservoir’s riparian zone [36]. Moreover, the deep water created by reservoir storage, often reaching hundreds of meters, limits the transmission of solar light and heat to the lower layers, leading to thermal and oxygen stratification [46,50]. Although almost all nutrients were stored at the bottom layer of reservoirs, especially under the summer stratification regime, the cold, dark bottom sediment environment still restricts the survival and reproduction of submerged aquatic plants [158,159]. Consequently, the application of phytoremediation techniques in high dam reservoirs is challenging.

In high dam reservoirs, traditional sediment capping techniques are often unsuitable due to the challenges of sediment removal and water reservoir management [11,12]. Periodic sediment dredging and reservoir discharge lead to the removal of surface sediment and its discharge downstream. The clean capping material is difficult to maintain on the reservoir bed, which undermines its ability to serve as an effective physical and chemical barrier [18]. Additionally, the significant volume of sediment required for a ~30 cm thick capping layer would greatly increase sediment accumulation in the reservoir, which is undesirable for hydropower development [46]. Therefore, in the context of high dam reservoirs, solidification/stabilization techniques maybe the most effective method for managing heavy metal pollution. Recently, the use of various types of stabilizing agents, such as nano-chlorapatite modified biochars, has been attempted for treating heavy metal pollution in reservoir sediments [139]. However, further evaluation of the practical heavy metal stabilization effects of these stabilizing agents in reservoir environments will be necessary in the future, along with an exploration of their potential impacts on the aquatic ecological environment.

6. Conclusions

This review examines the impact of global hydropower development and river damming on heavy metal transport. It outlines hydropower’s status, detailing how damming alters hydrology, physicochemistry, and ecology, thereby influencing heavy metal distribution in water, suspended sediments, and riverbed sediments. The key focus is on damming’s effects: promoting metal migration from water to suspended sediments, enhancing deposition in riverbed sediments, and triggering release during reservoir regulation. It summarizes in situ (stabilizing agents, capping, phytoremediation) and ex situ (dredging, chemical washing) remediation techniques, emphasizing in situ methods as practical for reservoirs. Overall, this review aims to provide insights for the sustainable management of river-reservoir systems affected by heavy metal pollution.

In conclusion, current research on heavy metal transport under damming has made progress, but several critical knowledge gaps remain. First, the cumulative effects of cascade hydropower systems on long-term heavy metal cycling are poorly understood, especially across large river basins. Second, the interaction between thermal-oxygen stratification in deep reservoirs and microbial-driven metal transformation is underexplored, which limits a comprehensive understanding of the biogeochemical mechanisms governing metal mobility. Third, field-scale validation of in situ remediation techniques is lacking, particularly regarding their adaptability to extreme environments such as high-altitude, nutrient-poor headwater reservoirs. To address these gaps, future studies should (1) investigate multi-reservoir cascades using isotopic tracers (e.g., Pb, Cd, or Sr isotopes) to trace heavy metal sources and migration pathways, combined with in situ sensors for real-time monitoring of hydrological parameters (e.g., flow velocity, turbidity) and dissolved metal contents, thereby clarifying cumulative metal transport patterns across large basins; (2) combine microbial metagenomics (to characterize functional microbial community composition and metabolic pathways) with geochemical analysis (e.g., sequential extraction for metal speciation determination) to unravel the mechanisms of microbe-heavy metal interactions under thermal-oxygen stratification in deep reservoirs; and (3) develop tailored in situ remediation technologies for extreme reservoir environments and conduct pilot-scale sediment capping tests using region-specific modified materials to validate their long-term stability, cost-effectiveness, and adaptability. Addressing these research gaps will improve pollution control and support sustainable hydropower development.