4.2.2.1 Indus Basin Project

The Indus Basin Project is the world’s largest irrigation system and includes some 26.3 Mha of cultivated lands within a dry region fed by one of the world’s largest rivers. The irrigation project was born from the Indus Water Treaty signed in 1960, which has been a model of trans-boundary cooperation between India and Pakistan for seven decades. To manage water supply to the irrigation district three large reservoirs were constructed (Chashma, Mangla, and Tarbela). Additional hydraulic infrastructure comprising the irrigation project includes twelve inter-river canals, forty-five canals extending 60,800 km, twenty-three barrages and siphon structures, communal watercourses, farm channels, and field ditches that extend some 1.6 million km to provide water to over 90,000 farmer-organized water courses (FAO, 2016). The irrigated lands provide some 95% of Pakistan’s food production. But the system is fragile, needing repairs, and under stress because of increasing competition from hydropower dams in the Himalayan headwaters of the Indus basin (Raza et al., Reference Raza, Naeem and Qadeer2019).

Despite the long-term effectiveness of the Indus Water Treaty, the race to build mega dams in the Indus basin headwaters are straining relations between India and Pakistan, as well as China (Laghari et al., Reference Laghari, Vanham and Rauch2012; Raza et al., Reference Raza, Naeem and Qadeer2019). India is building several large hydropower plants in the Sutlej, Beas, and Ravi River basins that will impound streamflow that drains to Pakistan (FAO, 2016). China is building dams on the upper Sutlej River in China, which then drains to India before eventually joining the Indus River in Pakistan as a major tributary. Pakistan is constructing the Diamer-Bhasha Dam (2028 completion), which at 272 m high will be the tallest concrete filled roller dam in the world and have a massive storage capacity of 10 km³. Pakistan is also constructing the Bunji Dam (190 m high) on the upper Indus River main-stem as well as five additional dams to form a North Indus River cascade.

Collectively, the new Indus River basin dams will reduce water and sediment flow to the largest irrigation project in the world, particularly in Pakistan’s Punjab and Sindh provinces, the latter of which includes much of the heavily degraded Indus delta (Qureshi, Reference Qureshi2011; FAO, 2016). Already some 95% of Indus River water is extracted for irrigation before reaching the coast, with zero flow in prolonged dry periods (FAO, 2011). This means that the extensive irrigation districts in the dry lower basin and delta will experience additional water stress (Qureshi, Reference Qureshi2011), with expected declines in riparian and delta fisheries (Laghari et al., Reference Laghari, Vanham and Rauch2012).

Overall, new dam construction in the Indus Himalayan headwaters is projected to result in a modest increase in river fragmentation (RFI from ~60% to ~75% over 2010–2030), but because the structures are mega dams the degree of impoundment will be staggering, with the RRI increasing from ~30% to ~180% over 2010–2030 (Grill et al., Reference Grill, Lehner, Lumsdon, MacDonald, Zarfl and Liermann2015).

4.3.1.1 United States

The wide physical geographic variability across the United States elucidates regional differentiation on the downstream impacts of dams to streamflow regime. And the greatest impacts occur along rivers with a high ratio of reservoir storage capacity to mean annual water yield (~dry regions). At a national scale, very large dams in the United States have reduced annual peak discharge pulses by 67% and low-flow discharge has increased by 52%. The ratio between the annual maximum and mean flow has decreased by 60% (Graf, Reference Graf2006). These hydrologic changes are epitomized along the middle Missouri River, for example, where the post-impoundment discharge during the spring high-flow period is significantly lower, a critical issue because of coinciding during periods of fish spawning (Pegg et al., Reference Pegg, Pierce and Roy2003).

An additional important streamflow alteration with ecological implications is changes to the timing (seasonality) of high- and low-flow periods. This often occurs because of dam management in relation to stakeholder interests (e.g., irrigation schedules, electricity demands, recreation, flood management). In such cases, critical high-flow pulses may be offset by weeks to months. Rather than coinciding with the “natural” wet period, higher flows occur during low-flow periods (Graf, Reference Graf2006). The most problematic aspect of changes to the timing of high- and low-flow periods occurs when they are out of accord with the environmental rhythms of the riparian corridor, such as seed germination, avian water fowl nesting, or the dependence of fisheries lifecycles for feeding and spawning (Graf, Reference Graf2002; NRC, 2005; Górski et al., Reference Górski, van den Bosch and van de Wolfshaar2012), as discussed below.

The reduced streamflow variability of US rivers is mainly consistent with international trends in changes to downstream streamflow regime following impoundment and dam closure, with variation due to basin-specific physical and stakeholder influences, particularly in support of agriculture.

4.3.1.2 Europe: Ebro and Volga Rivers

Extensive national and regional analyses of the impacts of dams on flow regimes similar to that conducted in the United States (i.e., Graf, Reference Graf2006) have not been conducted for Europe. It is well acknowledged, however, that streamflow regimes of European rivers are heavily impacted by dams. Indeed, 90% of Earth’s most obstructed rivers are located in Europe (Kornei, Reference Kornei2020). An extensive analysis of pre- and post-dam streamflow trends for the Ebro basin (85,530 km²) in Spain identified reduced flow variability (Batalla et al., Reference Batalla, Kondolf and Gomez2004). This was manifest by higher low flows due to reservoir outflow during dry periods (summer) to support irrigation while the storage of winter streamflow reduced maximum flow values. The two-year and ten-year flood magnitudes were reduced by over 30% on average for most gauging stations, including in the lower Ebro River (at Tortosa) downstream of the Mequinenza and Ribarroja reservoirs (Figure 4.12).

Figure 4.12.

Changes to flow regime for two large European rivers, before and after impoundment.

(A)

Reduction in annual maximum floods, lower Ebro River (at Tortosa) downstream of the Mequinenza and Ribarroja reservoirs.

(B)

Changes to lower Volga River discharge regime after dam closure over 1959–1960.

(Sources: Ebro River, Batalla, 2003; Volga River, Schmutz and Moog, 2018 with data from Górski et al., 2012.)

Like many large northern midlatitude rivers, the Volga (1,400,000 km²) serves as a major economic corridor. The Volga River is navigable for some 3,200 km through a series of eleven dams and canals that link northern ports along the White and Baltic to southern ports in the Caspian and Black Seas (Avakyan, Reference Avakyan1998; Middelkoop et al., Reference Middelkoop, Alabyan, Babich, Ivanov, Hudson and Middelkoop2015; Rodell et al., Reference Rodell, Famiglietti and Wiese2018). The Volga annually handles two-thirds of inland Russian navigation. Following impoundment of the lower Volga River behind the Volgograd Hydroelectric station in 1959 (at closure the largest hydroelectric dam in the world, remains largest in Europe), annual flow variability declined overall, including an increase in low flows by over ~1,000 m³/s (Figure 4.12B). The average maximum discharge reduced from 34,500 (m³/s) prior to dam closure (1959) to 26,800 (m³/s) after impoundment between 1959 and 1999 (Korotaev et al., Reference Korotaev, Ivanov and Sidorchuk2004). Of key importance to the hydrology of the lower Volga River is the vast amount of water withdrawn for industry, domestic use, and especially irrigation for agriculture, with the latter annually accounting for some 200 Mha of cultivated lands. Water consumption greatly increased following dam construction, increasing from 6 km³/yr pre-dam to 24 km³/yr post-dam, an amount that is about 10% of the water discharged into the Caspian Sea (Middelkoop et al., Reference Middelkoop, Alabyan, Babich, Ivanov, Hudson and Middelkoop2015). The Volga contributes 80% of the annual runoff supplied to the Caspian Sea, and annual water withdrawals for agriculture are contributing to Caspian Sea levels lowering (Rodell et al., Reference Rodell, Famiglietti and Wiese2018).

4.3.1.3 Yangtze and Three Gorges Dam

The streamflow regime downstream of the world’s largest dam, Three Gorges Dam (closure in 2003) along the middle Yangtze River has also changed downstream streamflow patterns (Zhang et al., Reference Zhang, Yuan, Han, Huang and Li2016; Tian et al., Reference Tian, Chang, Zhang, Wang, Wu and Jiang2019). Reduced flow variability extends some ~1,600 downstream, with reductions ranging from about 11% below the dam to 4% near the delta (Wang et al., Reference Wang, Shen, Gleason and Wada2013). These modest changes in river stage, however, mask hydrologic problems associated with management of Three Gorges Dam, in particular the abrupt up and down ramping and flow release schedule that produces frequent out-of-season flow variability. The flow release results in an oscillating pattern of drying out and backwater flooding of the massive floodplain lake system, including Poyang Lake (Zhang et al., Reference Zhang, Yuan, Han, Huang and Li2016). Not surprisingly, and with fair warning (i.e., Leopold, Reference Leopold1998), the downstream riparian environment has substantially degraded since closure of Three Gorges Dam, in association with sediment reduction and geomorphic adjustment (Zhang et al., Reference Zhang, Yuan, Han, Huang and Li2016; Cheng et al., Reference Cheng, Opperman, Tickner, Speed, Guo and Chen2018).

4.3.1.4 Nile, Aswan High Dam (Lake Nasser and Lake Nubia), and Water Politics

Because of its link to Egyptian antiquity, the historic Nile streamflow regime is perhaps the world’s most iconic coupled agro–hydro system, with irrigated agriculture extending along the entire Nile valley and across the delta plain (Figure 4.13). For over six millennia the Nile annually delivered nutrient-rich floodwater and silts to irrigated agricultural lands. But this abruptly ceased upon closure of the Aswan High Dam over 1965–1967 (Figure 4.14). The prior dynamic flow regime characterized by a seasonal fluctuation in river levels of 5–8 m has been greatly reduced to 1–2 m. As is common to downstream flow variability, Nile River low flows are moderately higher while the large annual flood pulses are greatly reduced, from ~8,000 m³/s to less than ~3,000 m³/s (Figure 4.15). A system of weirs downstream of Aswan High Dam maintains moderate flow stage levels to increase the river’s irrigation potential. Nile impoundment has greatly reduced annual runoff volumes, from 80 km³ to 30 km³ from pre-dam to post-dam, respectively (Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017). Although agricultural productivity increased with impoundment, the high demands on the irrigation system results in frequent water stress for some 95% of Egypt’s riparian population dependent upon the Nile River for 90% of its freshwater.

Figure 4.13.

The Nile extends ~1,000 km from Aswan High Dam (bottom of image) to the delta without tributary inputs. Riparian agriculture effectively covers the entire valley and delta and contrasts with the hyper-arid Sahara.

(Licensed by CC.)

Figure 4.14.

Lake Nasser and Aswan High Dam, and Nile River in Egypt. Lake Nasser formed with dam closure over 1965–1967. The reservoir covers 5,250 km² and extends for ~500 km, with 350 km as Lake Nasser in Egypt and 150 km as Lake Nubia in Sudan. Storage capacity for the reservoir is 162 billion m³. Reservoir sedimentation is measured regularly by bathymetric surveys along twenty-one fixed cross sections (Ahmed and Ismail, Reference Ahmed and Ismail2008). Scale: main axis of reservoir is ~3 km in width. North is to lower right of image.

(Source: NASA photo, April 12, 2015.)

Figure 4.15.

Streamflow variability of Nile River below Aswan Dam between 1950 and 1985, revealing impacts to annual flood pulse (higher low flows, much lower peak flows) after dam closure over 1965–1967.

(Figure Source: Vörösmarty and Sahagian, 2000.)

Today the main issue concerning Nile River management is located some 3,200 km upstream of Cairo and concerns impoundment of the Blue Nile by the Grand Ethiopian Renaissance Dam. As the largest reservoir in Africa, the Grand Ethiopian Renaissance Dam will store 1.5 times the average annual flow of the Blue Nile (Wheeler et al., Reference Wheeler, Basheer and Mekonnen2016). This will greatly increase the index of river impoundment (i.e., Grill et al., Reference Grill, Lehner, Lumsdon, MacDonald, Zarfl and Liermann2015) and, based on lessons learned from dam impacts on dryland rivers in the United States, the downstream riparian zone will be substantially degraded because of reduced flow variability (e.g., Graf, Reference Graf2006).

For well over a decade, imminent impoundment of the Blue Nile in Ethiopia has increased water stress and political tensions between Egypt and Ethiopia. Aside from the longer-term management scheme of the dam and reservoir – such as changes in water quantity, quality, and the outflow release schedule – a pressing issue receiving considerable international attention is the period of reservoir infilling (Wheeler et al., Reference Wheeler, Basheer and Mekonnen2016; El-Nashar and Elyamany, Reference El-Nashar and Elyamany2018). If the reservoir is rapidly infilled, it requires that the reservoir impounds all upstream Blue Nile waters (zero outflow). Even this extreme measure will require about three years to fill the reservoir to a level where hydropower turbines can be activated. And this is dependent upon the rainfall patterns over the Ethiopian highlands (El-Nashar and Elyamany, Reference El-Nashar and Elyamany2018). Such a rapid rate of reservoir infilling would effectively desiccate downstream riparian lands in Sudan, upstream of Egypt’s Aswan High Dam. International agreements between Nile basin nations with regard to water withdrawals have been especially ineffective but are particularly critical in view of imminent closure of the Grand Ethiopian Renaissance Dam (Wheeler et al., Reference Wheeler, Basheer and Mekonnen2016).

4.3.2.1 Global Perspective

The global proliferation of dams means that copious amounts of Earth’s riverine sediments are trapped in upstream reservoirs, small and large (e.g., Morris and Fan, Reference Morris and Fan1998; Walling, Reference Walling2006; UNESCO-IHP, 2011; Tockner et al., Reference Tockner, Bernhardt, Koska, Zarfl and Hüttl2016). Trap efficiencies of main-stem dams along large rivers tend to be very high (Table 4.4), with many reservoirs trapping greater than 80% of upstream sediment flux (Vörösmarty et al., Reference Vörösmarty, Meybeck, Fekete, Sharma, Green and Syvitski2003). Many US dams, including large dams in the Mississippi basin, have trap efficiencies greater than 95%. The large main-stem Colorado dams trap up to 100% of upstream sediment loads (except when intentionally flushed), as does the Nile River upstream of the Aswan High Dam (Williams and Wolman, Reference Williams and Wolman1984).

Table 4.4Impacts of reservoir sediment trapping along several large rivers

Source: Walling (Reference Walling2006), with data reported from Vörösmarty et al. (Reference Vörösmarty, Meybeck, Fekete, Sharma, Green and Syvitski2003).

Fluvial sediment delivery to the coast has precipitously declined since the mid-1900s (Syvitski et al., Reference Syvitski, Vörösmarty, Kettner and Green2005; Walling, Reference Walling2006; UNESCO-IHP, 2011). Globally, some 12.6 GT of sediment loads are annually discharged into the oceans (Table 4.5). While estimates vary widely, the global decline in sediment flux ranges from 16% to 66% of annual riverine sediment loads (Syvitski et al., Reference Syvitski, Vörösmarty, Kettner and Green2005; Walling, Reference Walling2006). Much of the decline in sediment flux is attributed to the impact of dams, although improved land management, channel engineering, and climate change also reduces sediment loads (Batalla et al., Reference Batalla, Vericat and Martínez2006; Wang et al., Reference Wang, Yang, Saito, Liu and Xiaoxia2007; Walling, Reference Walling2008b; Meade and Moody, Reference Meade and Moody2010; Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017). Because dam construction coincides with different periods of human activities and climatic fluctuation, changes to sediment loads after dam closure can vary greatly from river basin to river basin (Table 4.6).

Table 4.5Comparison of two estimates of the global fluvial sediment budget and its modification by human activity

Source: Walling (Reference Walling2006).

Table 4.6Comparison of sediment yields and runoff before and after dam impoundment for selected rivers

Data: Liu et al. (Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017).

Reservoir sediment storage has diverse implications to fluvial environments. This includes upstream, within, and downstream of the reservoir (Table 4.7). Reservoir sedimentation can result in upstream flooding and reduced navigation, and aquatic habitat degradation. Problems associated with sediment storage in reservoirs includes reduced hydropower potential because of reduced storage capacity, storage of contaminated sediments, and reservoir stratification impacting aquatic habitat within the reservoir and downstream (discussed further, below). The high sediment trap efficiency of many dams substantially reduces downstream fluvial sediment loads, with consequences to geomorphic adjustments and riparian habitat. In addition to the quantity of sediment being reduced, the quality of sediment is often also reduced. In combination with the altered flow regime fine bed material is removed, resulting in downstream channel bed material becoming coarser, and in some cases leading to channel bed armoring.

Table 4.7Environmental and functional issues related to dam and reservoir sediment storage

Sources: Morris and Fan (Reference Morris and Fan1998), Scheueklein (Reference Scheueklein1990), Randle et al. (Reference Randle, Morris and Whelan2019), Winton et al. (Reference Winton, Calamita and Wehrli2019), Morris (Reference Morris2020), and author.

4.3.2.2 Sediment Decline to Nile, Link to Agriculture

For millennia, rains in the Ethiopian Highlands generated runoff that transported some 95% of the Nile River’s sediment load. Sediment was principally supplied to the main-stem Nile by the Blue Nile (324,530 km²) and Atbara (166,875 km²) Rivers, both of which enter the main-stem Nile upstream of Aswan High Dam. But after impoundment behind the Aswan High Dam all upstream Nile sediment is stored in Lake Nasser (Ahmed and Ismail, Reference Ahmed and Ismail2008; Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017). Suspended silt concentrations averaged 3,800 mg/L between 1929 and 1963 but abruptly declined to only 129 mg/L by 1965, two years after the start of infilling Lake Nassar reservoir (Raslan and Salama, Reference Bakkenist and Flos2015). The average annual sediment load in the pre-dam period was 120 million tons, but was reduced to 0.2 million tons after dam construction (Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017). Further downstream, sediment loads in the Nile only marginally recover, as there are no appreciable tributaries for about the final 1,000 km (Figure 4.16). A minor amount of sediment is supplied from aeolian inputs, bank erosion and channel bed scour. Low flow structures (weirs) control stage levels in support of navigation and floodplain irrigation (Morris and Fan, Reference Morris and Fan1998; Ahmed and Ismail, Reference Ahmed and Ismail2008). At the coast the Nile discharges very little sediment (~5 × 10⁶ tons/yr) to the Mediterranean, with the substantial reduction in sediment loads contributing to coastal erosion and retreat of the Nile Delta (Stanley and Warne, Reference Stanley and Warne1998).

Figure 4.16.

Downstream changes to suspended sediment loads along the Nile River before and after closure of Aswan High Dam in 1964. Suspended sediment comprises 30% clay, 40% silt, and 30% fine sand (Ahmed and Ismail, Reference Ahmed and Ismail2008).

(Source: Morris and Fan, 1998.)

Impacts to the Nile sediment and streamflow regime caused by the Aswan High Dam aren’t merely significant, they’re symbolic. The lower Nile is the premier example of a millennia-old, intensive floodplain-agricultural coupled system, especially dependent upon annual floodwaters and sediment for irrigation and fertility. As an early hearth of irrigation science, Nile farmers utilized an ingenious system of sluice gates, water wheels, dams, and buckets to maximize the potential of furrow-and-field irrigation methods (Butzer, Reference Butzer1976). But the iconic floodwater irrigation system was all but eliminated with construction of the Aswan High Dam, and initially reduced downstream productivity. Crop yields have increased since the dam was completed, but at a high cost. The Nile delta has continued to degrade because of a lack of fluvial sediment (Becker and Sultan, Reference Becker and Sultan2009; Stanley and Clemente, Reference Stanley and Clemente2014), and Nile delta agriculture now requires significantly larger “hard engineering” structures, including larger canals, ditches, and larger mechanical pumps for irrigation and drainage (Molle, Reference Molle2018). Additionally, the Nile agricultural system is dependent upon high amounts of chemical fertilizer to maintain high agricultural yields, an issue that results in interesting and unintended impacts to coastal fisheries production (Nixon, Reference Nixon2003), as noted further in text. Of course, the artificially high water table is problematic with regard to the annual deficit in rainfall (165 mm/yr) relative to evapotranspiration (1,500 mm/yr), and results in land degradation through soil salinization (Ahmed and Ismail, Reference Ahmed and Ismail2008; Molle, Reference Molle2018).

4.3.2.3 Sediment Decline to Yangtze and Huanghe Rivers

Some of the large rivers in Asia have more recently experienced sharp reductions in sediment loads following dam construction, and the changes continue to rapidly unfold because of many new dam construction projects (Syvitski et al., Reference Syvitski, Vörösmarty, Kettner and Green2005; van Binh et al., Reference van Binh, Kantoush and Sumi2019). This is particularly true of the Yangtze and Huanghe Rivers in China. Suspended sediment loads along the Yangtze River downstream of Three Gorges Dam at the lowermost sediment station (Datong, 565 km upstream of the delta) were already in decline. In the 1950s and 1960s, suspended sediment load averaged 507 Mt/yr but had declined in the period before dam construction to 320 Mt/yr (1993–2003). After closure of Three Gorges Dam (TGD), the sediment loads between 2003 and 2012 steeply dropped (Figure 4.17), to 145 Mt/yr. Between 2003 and 2012, an astonishing 182 Mt/yr (80% of the total sediment load) was trapped behind Three Gorges Dam (Yang, et al., Reference Yang, Xu, Milliman, Yang and Wu2015). Between 2002 and 2008, the median particle size (d₅₀) of the channel bed material increased from 0.36 mm to 25 mm at Yichang (44 km downstream of TGD) but only increased from 0.18 mm to 0.19 mm at Chenglinji (420 km downstream of TGD) (Zheng et al., Reference Zheng, Xu, Cheng, Wang, Xu and Wu2018). Between these two stations (at 195 km downstream of TGD), the d₅₀ of bed material increased from 0.19 mm (pre–TGD) to 0.251 mm by 2010, seven years after closure (Zhang et al., Reference Zhang, Yuan, Han, Huang and Li2016).

Figure 4.17.

Comparison of annual discharge and suspended sediment loads before and after impoundment of the Yangtze River by Three Gorges Dam: Pre-dam (1993–2002) and post-dam (2003–2012). Datong is 565 km upstream of the river mouth.

(Source: Yang et al., 2015.)

The Huanghe River is also referred to as the Yellow River because of its ochre colored sediment that derives from erosion of the Chinese Loess Plateau. The Huanghe previously transported 1.08 Gt of sediment per year to the coast (at Lijin, 40 km from delta) – Earth’s highest annual sediment load. By 2005 this had declined to 0.15 Gt/yr, which is only 14% of its historic sediment load (Wang et al., Reference Wang, Yang, Saito, Liu and Xiaoxia2007).

A key reason for the decline in sediment loads of the Huanghe River is construction of seven main-stem dams between 1968 and 1998 (Table 4.8). It should be noted, however, that in addition to main-stem dam emplacement several other explanations are also responsible for the large decline in sediment loads over the past several decades, for both the Huanghe and Yangtze Rivers. These include improved land management (noted in Chapter 2), increased water withdrawals for agriculture and industry, and reduced precipitation due to climate change (Walling, Reference Walling2006; Wang et al., Reference Wang, Yang, Saito, Liu and Xiaoxia2007; Yang et al., Reference Yang, Xu, Milliman, Yang and Wu2015; Shi et al., Reference Shi, Hu, Deng, Tian, Wieprecht, Haun, Weber, Noack and Terheidenet2017). For the Huanghe, the impact of main-stem dam construction and improved management of the Loess Plateau has been so effective at reducing downstream sediment loads that downstream channel engineering measures were implemented to manage channel bed incision. Between the 1960s and 2000, soil conservation in the Chinese Loess Plateau reduced sediment inputs into the Huanghe River by some ~300 × 10⁶ tons/yr. On the Yangtze, 65% of the reduction in downstream sediment loads since 2003 is attributed to closure of Three Gorges Dam, 5–14% is attributed to climate change (reduced precipitation), and the remaining cause of sediment load reduction is attributed to upstream dams and soil conservation (Chen et al., Reference Chen, Wei, Fu and Lu2007; Yang et al., Reference Yang, Xu, Milliman, Yang and Wu2015).

Table 4.8Characteristics of main-stem reservoirs along the Huanghe River, China

Source: Shi et al. (Reference Shi, Hu, Deng, Tian, Wieprecht, Haun, Weber, Noack and Terheidenet2017).

4.3.2.4 Sediment Decline Following Mekong Dam Construction

The Mekong delta receives much less sediment than before construction of a cascade of dams along the upper Mekong (Lancang) in China, which started with closure of Manwan Dam in 1993. The most recent analysis of “before and after” comparison of sediment loads shows a large sediment flux reduction to the Mekong delta, by some 74%. The pre-dam (before 1993) sediment flux to the Mekong delta was 166.7 million tons/yr (±33.3) but has now declined to 43.1 million tons/yr for 2012–2015 (van Binh et al., Reference van Binh, Kantoush and Sumi2019), which varies by ±20 million tons per year depending on whether ENSO is in a negative or positive phase, respectively (Ha et al., Reference Ha, Ouillon and van Vinh2018).

Closure of the Manwan Dam on the upper Mekong main-stem abruptly reduced downstream suspended sediment loads by >60% at Gajiu, 2 km below the dam, with modest sediment recovery further downstream (Fu et al., Reference Fu, He and Lu2008). Between 1993 and 2003, 26.9–28.5 million tons/yr were trapped in the Manwan reservoir, resulting in reservoir sediment storage of 295.9–313.5 million tons over an eleven-year period (Fu et al., Reference Fu, He and Lu2008). The large amount of reservoir sedimentation is concerning, as 21.5–22.8% of the total storage capacity of the Manwan reservoir was lost over an eleven-year period following impoundment. As a whole, the trap efficiency of six large main-stem dams along the upper Mekong in China ranges from 61% to 92% (Figure 4.18). As regards downstream sediment loss this is alarming when considering that some thirteen additional main-stem dams are planned or under construction upstream, toward the basin headwaters (Kummu and Varis, Reference Kummu and Varis2007; Fan et al., Reference Fan, Hea and Wang2015) (Figure 4.19). And already thirty-seven more dams exist in the lower basin in Laos, Cambodia, Thailand, and Vietnam.

Figure 4.18.

The “stair-step” profile of a ~1,000 km segment of the upper Mekong (Lancang) River in China, upstream of Laos border, including trap efficiency (%), hydraulic head, dam height, and year of closure. *Upstream of this reach, seven large main-stem dams are under construction, including Miaowei (140 m high), Dahuaqiao (106 m), Huangdeng (202 m), Tuoba (158 m), Lidi (74 m), Wunonglong (136.5 m), and Gushui (220 m); and further upstream, six additional main-stem hydroelectric dams are planned or in construction, extending the stair-step profile toward the Himalayan headwaters.

(Source: InternationalRivers.org Fact Sheet on upper Mekong (Lancang) Dams, May 2013, accessed April 12, 2020. Figure source: Kummu and Varis, 2007, with author modifications and updates.)

Figure 4.19.

Location of large dams (> 15-m height) in the Mekong basin, including completed, under construction, and planned. The border between the Upper Mekong (Lancang) and Lower Mekong divide is indicated. The “3S” river basins refer to the Sekong, Sesan, and Srepok Rivers, which join to form the largest tributary in the lower Mekong.

(Author figure. Data sources include Mekong River Commission and Räsänen et al., 2017 (for large dams).)

Detecting changes to the Mekong delta sediment flux highlights the importance of having continuous sediment records before and after impoundment, having sediment gauging stations in key upstream and downstream locations within the drainage basin, and having access to the data (Walling, Reference Walling2008a, Reference Walling and Campbell2009). A study by van Binh et al. (Reference van Binh, Kantoush and Sumi2019) provides the longest period of analysis – spanning five decades – that envelops the pre- and post-dam periods (1961–2015). Crucially, the analysis utilizes sediment load data for years 1965–2003 at the Jiuzhou station (China), above the upper Mekong cascade of dams.

The post-dam period for the Mekong River is a story that continues to unfold because of construction and planning of numerous large dams (Figure 4.19) throughout the basin (Walling, Reference Walling2008a,Reference Wallingb; Pokhrel et al., Reference Pokhrel, Burbano, Roush, Kang, Sridhar and Hyndman2018). Because of the large amount of drainage area in the lower Mekong, there is concern over the ongoing and future development of forty-two dams in the Se San, Srepok, and Se Kon Rivers in the lower basin, which join together to form the largest tributary to the Mekong. If the 133 dams planned for the Mekong are constructed only 4% of the Mekong River sediment load will be discharged to the delta (Kondolf et al., Reference Kondolf, Gao and Annandale2014a, Reference Kondolf, Rubin and Minear2014b; van Binh et al., Reference van Binh, Kantoush and Sumi2019), which would be devastating to the deltaic geomorphology and associated aquatic environments (Piman and Manish, Reference Piman and Manish2017).

4.3.2.5 Sediment Decline for Several European Rivers: Rhine, Ebro, Volga

The sediment loads of European rivers have been impacted by humans for centuries and millennia. Several rivers are representative of the impacts of dams to European rivers, including the Ebro, Volga, and Rhine Rivers.

A longitudinal profile of the Rhine River basin reveals the artificial stair-step pattern in the Upper and High Rhine because of intensive impoundment. This includes the Iffezheim Dam, the lowermost and last dam to be emplaced along the main-stem Rhine River (Figure 4.20). Suspended sediment loads along the Rhine River (185,000 km²) has increased and decreased with episodes of land cover change and hydraulic engineering (Vollmer and Goelz, Reference Vollmer and Goelz2006; Spreafico and Lehmann, Reference Spreafico and Lehmann2009; Middelkoop et al., Reference Middelkoop, Erkens and van der Perk2010; van der Perk, Reference van der Perk, Sutaria, Middelkoop, Stouthamer, Middelkoop, Kleinhans, van der Perk and Straatsma2019). The more recent episode, since about the 1950s, reveals about a 70% decline, which is dramatic for a river that had already been intensively impacted by human activities. Annual suspended sediment loads declined from 4 × 10⁶ tons to 1.2 × 10⁶ tons by 2016, and have remained relatively stable over about the past ten years (van der Perk, Reference van der Perk, Sutaria, Middelkoop, Stouthamer, Middelkoop, Kleinhans, van der Perk and Straatsma2019). The primary source of the decline since the early 1950s is likely the completion of two main-stem dams in the Upper Rhine, specifically Gambsheim and Iffezheim dams (Figure 4.20) completed in 1974 and 1977, respectively. As a whole there are twenty-one main-stem dams on the Upper and High Rhine. The sediment starvation problem along the Rhine has resulted in channel bed incision along large reaches of the main-stem Rhine in the delta and alluvial valley (Quick et al., Reference Quick, König, Baulig, Schriever and Vollmer2020) that threatens navigation and associated infrastructure, requiring innovative sediment management approaches (see Section 8.2.3).

Figure 4.20.

Longitudinal profile of the Rhine River, including geomorphic province, major tributaries, and sediment management operations. The stair-step pattern in the Upper and High Rhine is due to intensive impoundment. Gambsheim hydroelectric dam is located at 309-km, 25 km upstream of Iffezheim dam.

(After Götz, 2008 and Frings et al., 2019.)

The sediment flux of the lower Ebro River (85,530 km²) was historically high, annually discharging 20–30 million tons/yr into the Mediterranean (late nineteenth century). But the Ebro suspended sediment loads abruptly declined to 3.3 million tons/yr following a period of dam building in the early 1960s. By the early 2000s, and ~299 dams later, the sediment load transported to the delta had declined to a meager 0.29 million tons/yr – less than 2% of pre-dam sediment loads (Rovira and Ibàñez, Reference Rovira and Ibàñez2007). The dramatic reduction in Ebro River sediment load has had substantial adverse consequences to its delta, and ground subsidence is of particular concern (Tena and Batalla, Reference Tena and Batalla2013). About 45% of the emergent Ebro delta will be drowned by 2100 because of subsidence and rising sea levels. The strategy to mitigate against higher sea levels includes building new lands by sediment diversion structures within the delta and managing reservoir sediment by flushing operations to mobilize stored reservoir deposits for transport through a cascade of reservoirs (Rovira and Ibàñez, Reference Rovira and Ibàñez2007; Tena and Batalla, Reference Tena and Batalla2013).

Impoundment of the lower Volga River behind the Volgograd Dam trapped sediments and resulted in a large decline in annual suspended sediment loads. Prior to dam closure (1934–1953), the suspended sediment loads of the lower Volga River were 12–18.5 million tons/yr, which declined to 7.4 million tons/yr in the post-dam period (1961–1982). The annual suspended sediment load at the delta apex substantially declined from the pre-dam to the post-dam period, being 26 million tons/yr and 7.9 million tons/yr, respectively (Middelkoop et al., Reference Middelkoop, Alabyan, Babich, Ivanov, Hudson and Middelkoop2015).

4.3.2.6 Sediment Decline to Missouri and Mississippi Rivers

The fact that upstream dam efficiencies are so high and that a lot of fluvial sediment still makes it to the coast implies that sediment loads somewhat recover (Williams and Wolman, Reference Williams and Wolman1984; Phillips et al., Reference Phillips, Slattery and Musselman2004). This occurs largely because of downstream channel scour and erosion (discussed subsequently) and because impacts of upstream dam trapping can be buffered by sedimentary inputs from downstream tributaries. The downstream sediment load recovery is often less than 50% of upstream sediment loads.

Following closure of the downstream-most main-stem dam in 1955 on the Missouri River, Gavins Point Dam in North Dakota and South Dakota (Figure 4.21), downstream suspended sediment loads immediately declined by 99%. At Hermann, Missouri, a distance of 1,047 km downstream of Gavins Point Dam, suspended sediment loads in the post-dam period (1957–1980) only recovered to ~30% of pre-dam amounts (Figure 4.22), which includes sediment inputs from the Platte River basin (10.1 × 10⁶ tons/yr) (Heimann, Reference Heimann2016). The precipitous decline in suspended sediment loads is detected further downstream in the middle Mississippi River (at St. Louis, Missouri), and all the way downstream to the Mississippi delta (Figure 4.23).

Figure 4.21.

Gavins Point Dam and Lewis and Clarke Lake, the lowermost dam on the Missouri River. The structure is an embankment dam of earthen and chalk-fill materials. Dimensions: closure in 1955, upstream drainage: 723,825 km², 23 m high, 2,650 m long, total reservoir capacity is 606,873 m³, surface area of 12,700 ha, depth of 14 m, and maximum length of 40 km.

(Photo and data source: U.S. Army Corps of Engineers.)

Figure 4.22.

Historic changes to suspended sediment loads along the Missouri and Mississippi Rivers in relation to impoundment by large main-stem dams.

(Source: Alexander et al., 2012.)

Figure 4.23.

Ratio of pre-dam and post-dam suspended sediment loads along the Missouri River downstream from Gavins Point Dam (closure in 1955) in South Dakota to below the confluence with the Mississippi River at St. Louis, Missouri. Stations and pre- and post-year dataset include Yankton (1940–1952, 1957–1969), Omaha (1940–1952, 1957–1973), St. Joseph, Kansas City, and Hermann (1949–1952, 1957–1976) located 8, 314, 584, 716, 1,147 km downstream from Gavins Point dam.

(Source: Williams and Wolman, 1984.)

As the case with many larger rivers, the timing of dam construction in the Mississippi basin occurred while other forms of hydraulic engineering and land cover change were also occurring, which further reduced downstream suspended sediment loads (Figure 4.24). Additional factors that reduced downstream suspended sediment loads in the Mississippi include improved land management in the upper basin (as with the Huanghe and Yangtze Rivers) and channel engineering in the lower basin. As channel bank erosion was previously a source of sediment for the lower Mississippi River, hydraulic engineering works such as groyens (wing dikes) and concrete revetment further reduce downstream sediment loads (Kesel et al., Reference Kesel, Yodis and McCraw1992; Kesel, Reference Kesel2003; Meade and Moody, Reference Meade and Moody2010).

Figure 4.24.

Historic change to suspended sediment loads for the lower Mississippi River at Tarbert Landing, Louisiana. Trend lines coincide with years 1950–1967 (−15 million tons/yr) and 1968–2006 (−1.1 million tons/yr). Cumulative distance (km) for revetment and wing dike (groyne) construction included for USACE Memphis and Vicksburg Districts.

(Source: Meade and Moody, 2010.)

4.3.3.1 Downstream Propagation of Channel Degradation

Spatially, along large rivers, channel bed degradation following impoundment may extend for tens to hundreds of kilometers downstream of the dam (Ma et al., Reference Ma, Huang, Nanson, Li and Yao2012; Smith et al., Reference Smith, Morozova, Pérez-Arlucea and Gibling2016; Smith and Mohrig, Reference Smith and Mohrig2017; Quick et al., Reference Quick, König, Baulig, Schriever and Vollmer2020). The amount of incision usually declines with downstream distance, being dependent upon the amount of change to the sediment and streamflow regime, the location and character of downstream tributary inputs, as well as the sedimentology and lithology of channel bed and bank material, valley slope, and human impacts (Brandt, Reference Brandt2000; Chin et al., Reference Chin, Harris, Trice and Given2002; Phillips et al., Reference Phillips, Slattery and Musselman2004; Smith et al., Reference Smith, Morozova, Pérez-Arlucea and Gibling2016; Lai et al., Reference Lai, Yin and Finlayson2017).

The closure of Three Gorges Dam in 2003 was a watershed moment in environmental sciences. Because of being the world’s largest dam on a very large river it has provided riparian sciences an unprecedented opportunity to examine the impacts of dam construction on a major river across a variety of fields. After closure of Three Gorges Dam, the Yangtze River abruptly transitioned from a depositional phase into an erosional phase. From the mid-1950s to mid-1980s the rate of channel bed deposition was about +90 Mt/yr of aggradation, but this reversed to about −50 Mt/yr of erosion following closure of Three Gorges Dam. Annual rates of average channel bed incision (estimated from low-flow stage data) between 2004 and 2012 decreased downstream, from 10.2 to 3.4 cm/yr at Jingjiang (50–420 km downstream TGD), 4.9 to 2.5 cm/yr in the middle reach (420–910 km downstream TGD), to 1.7 to 0.9 cm/yr in the lower reach (910–1,550 km downstream TGD) (Wang et al., Reference Wang, Shen, Gleason and Wada2013). Averages across the channel bed mask local variability, and along a 70-km long segment of the middle Yangtze (Shashi reach, ~190 km downstream of TGD), channel thalweg incision averaged 2 m between 2002 and 2010, with a maximum thalweg incision of 8 m (Zhang et al., Reference Zhang, Yuan, Han, Huang and Li2016). Channel scour moved progressively downstream, and it effectively ceases where the channel slope abruptly decreases. Sediment starvation is detected within the delta, manifest as reduced coastal marsh accretion and net erosion of the subaqueous delta front (Yang et al., Reference Yang, Milliman, Li and Xu2011; Yang et al., Reference Yang, Xu, Milliman, Yang and Wu2015; Zheng et al., Reference Zheng, Xu, Cheng, Wang, Xu and Wu2018). The well-documented case of the Yangtze River and Three Gorges Dam illustrates that following impoundment the geomorphology of large rivers can rapidly adjust.

Coastal draining rivers in Texas exhibit very different downstream responses to impoundment and illustrate the importance of drainage basin controls on the style of post-dam geomorphic adjustment (Williams and Wolman, Reference Williams and Wolman1984). Even within a region as homogeneous as the Texas coastal plain, fluvial geomorphic responses to dam impoundment can vary widely, complicating the development of general statements regarding the impacts of dams to rivers (Phillips, Reference Phillips2003).

From east to west Texas, the downstream riparian impacts of dam closure vary greatly, emphasizing the importance of streamflow variability to geomorphic response. In contrast to considerable downstream incision of the lower Trinity River (Phillips et al., Reference Phillips, Slattery and Musselman2005; Smith et al., Reference Smith, Morozova, Pérez-Arlucea and Gibling2016), the Sabine River along the Texas–Louisiana border revealed very little geomorphic adjustment to installation of Toledo Bend Reservoir in 1967. Sediment transport to the coastal zone continues, bars continue to accrete, and bends continue to migrate. This is likely because of sediment decoupling of the upper and lower reaches of larger basins, which buffers the downstream impacts of dams (Phillips, Reference Phillips2003). Additionally, the dam is a flow-over structure with little change to hydrologic regime (Phillips, Reference Phillips2003; Heitmuller and Greene, 2009). But to the west, along the lower Rio Grande/Bravo at the Texas–Mexico border, the impact of dam closure has been the opposite, and geomorphic and hydrologic processes have effectively been arrested.

The natural high-flow variability of the Rio Grande/Bravo along the US–Mexico border made the river more sensitive to change after impoundment (e.g., Graf, Reference Graf2006). A series of main-stem dams, beginning in 1916 with Elephant Butte dam in southern New Mexico, tremendously altered a once dynamic riparian environment that drains 557,722 km² of southwestern North America. Along the lower Rio Grande/Bravo, Falcon Dam and Reservoir (443 km upstream of Gulf of Mexico) was closed in 1953. This greatly reduced downstream stream power such that persistent sediment transport no longer occurs. The median and maximum recorded discharge in the pre-dam period (1934–1951) was 48.7 m³/s and 872 m³/s, respectively. But in the post-dam period (1954–2011) median and maximum discharge greatly reduced to 5.6 m³/s and 459 m³/s, respectively (Swartz et al., Reference Berkowitz, Johnson and Price2020). This is significant because suspended sand in the lower Rio Grande/Bravo is not transported until a discharge of ~6 m³/s (at median discharge). The majority of suspended sediment transport occurs by relatively infrequent events, specifically at discharge magnitudes ranging between 300 m³/s and 400 m³/s that have a flow duration of <5% (Hudson and Mossa, Reference Hudson and Mossa1997). Rates of meander bend migration declined from 11.6 m/yr to 0.9 m/yr between the pre-dam and post-dam periods, respectively (Swartz et al., Reference Berkowitz, Johnson and Price2020). The once great dynamic lower Rio Grande/Bravo is effectively locked in place.

Flow impoundment of the Rio Grande at Elephant Butte Reservoir (New Mexico) and downstream water withdrawals for irrigation have reduced streamflow to such a degree that aggradation, rather than incision, is the dominant form of geomorphic adjustment (Figure 4.27). At Presidio, Texas/Ojinaga, Chihuahua, Mexico, the Rio Grande/Bravo channel underwent some ~3 m of channel bed aggradation between 1933 and 1974, with about 20 m of channel narrowing. The reduced streamflow results in aggradation from smaller valley-side downstream tributaries because of a lack of flow competence of the main-stem Rio Grande/Bravo (Everitt, Reference Everitt1993; Collier et al., Reference Collier, Webb and Schmidt1996).

Figure 4.27.

Changes to Rio Grande/Bravo at Presidio, Texas/Ojinaga, Chihuahua.

(A)

Deposition between the levees (dikes) resulting in embanked floodplain aggradation.

(B)

+3 m of channel bed aggradation and narrowing due to reduced flow conveyance caused by upstream impoundment and water withdrawal.

(Source: Collier et al., 1996, data from Everitt, 1993.)

4.3.4.1 Terrestrialization of Riparian Environments

Reduced hydrologic variability and channel stabilization following upstream flow impoundment drives downstream riparian environmental change. Woody vegetation encroachment over formerly active channel bars and siltation of sloughs and backwater areas directly reduces critical aquatic habitat linked to specific biologic functions, such as fish feeding and spawning. Thus, strongly related to hydrologic, sedimentary, and geomorphologic changes noted earlier are changes to riparian ecosystems, including the types of flora and fauna associated with floodplain and channels (Ligon et al., Reference Ligon, Dietrich and Trush1995; Grams and Schmidt, Reference Grams and Schmidt2002; Schumm, Reference Schumm and Gupta2007).

For a large range of regulated rivers in the United States, a comparison with upstream unregulated reaches found that downstream regulated reaches have low-flow channels that are 32% wider and high-flow channels that are 50% narrower (Graf, Reference Graf2006). As relates to the riparian zone, floodplains of regulated rivers are 79% less active than upstream river reaches not influenced by dams. The functional portion of the river, which includes channel bar surfaces and floodplain environments in contact with streamflow at different frequencies and durations (Table 4.9), is 72% smaller for impounded reaches than upstream reaches not influenced by dams. And, in agreement with fragmentation indices of streamflow disruption, rivers within semiarid and arid regions with high-flow variability are especially heavily impacted because woody vegetation takes advantage of reduced flood pulse disturbances (Johnson, Reference Johnson1994; Marston et al., Reference Marston, Mills, Wrazien, Bassett and Splinter2005; Graf, Reference Graf2006). This scenario played out along the upper Snake River (4,510 km²) in Wyoming and Idaho (United States) following the closure of Lake Jackson Dam in 1906 (Marston et al., Reference Marston, Mills, Wrazien, Bassett and Splinter2005). In the case of the Snake River, species-poor forest vegetation assemblages, including Blue spruce, Cottonwood, and other mixed forest varieties degraded the riparian environment by replacing a rich mosaic of willow-alder and shrub-swampland and dynamic unvegetated surfaces that supported a diverse riparian ecosystem. Growth of woody vegetation reduced local channel avulsions and associated structural riparian diversity (geodiversity), further degrading lateral hydrologic connectivity. In the case of the upper Snake River, this degraded the native fishery habitat by reducing the spawning area for the Snake River fine-spotted cutthroat trout (Oncorhynchus clarkii).

Table 4.9Functional fluvial geomorphic surfaces and ecological relevance

Frequency definitions for the modern river adjustment to the contemporary geomorphic regime. Specific frequencies vary regionally.

From Graf (Reference Graf2006).

4.3.4.2 Changes to the Platte River, United States

The Platte River drains 219,900 km² of the US Midwest and Rocky Mountains and is a prime example of a large braided river impacted by impoundment. The Platte is a major sediment source to the Missouri River, and annually discharges 10.1 million tons of suspended sediment into the Missouri River at Plattsmouth, below Omaha, Nebraska (Heimann, Reference Heimann2016). Its dynamically adjusting channel and wetlands within a semiarid environment represent a crucial recharge station to migratory birds, including federally endangered Sandhill Cranes (Antigone canadensis). And the abundant low sandy channel bars are prime nesting habitat for interior least terns (Sterna antillarum) and piping plovers (Charadrius melodus), of concern because of historic degradation to this dynamic habitat (NRC, 2005).

The Platte River channel has considerably narrowed since the mid-1800s, which is primarily attributed to flow impoundment and water withdrawals for irrigation (Williams, Reference Williams1978b; Johnson, Reference Johnson1994; Schumm, Reference Schumm and Gupta2007). Between 1880 and 1995, average channel width decreased from ~1,250 m to ~300 m (Figure 4.28). The main culprit is the 1941 closure of Kingsley Dam (and Lake McConaughy reservoir), the world’s second largest hydraulic fill dam. And before completion of the reservoir the Platte River had experienced considerable riparian change because of irrigation withdrawals for agriculture, and closure of dams in Wyoming in the early 1900s (Johnson, Reference Johnson1994). This initiated a dramatic change to the riparian environment, as the 1938 channel width was about half its 1880 width (Williams, Reference Williams1978b; Schumm, Reference Schumm and Gupta2007). The trend in narrowing and stabilization continued until the mid-1990s. For the Platte River the increase in low flows is particularly problematic, and results in much of the formerly sandy braid plain and active channel bars being terrestrialized by woody vegetation.

Figure 4.28.

Changes to average channel width (m) along Platte River between Brady and Phillips, NE between 1880 and 1995. Kingsly dam is 118 km upstream of Brady Reservoir.

(Source: Schumm, 2007.)

Changes to the Platte River riparian corridor vividly illustrate the importance of maintaining a historic flow regime with regard to functional geomorphic surfaces (Table 4.9). Specifically, this includes flow variability (with higher high flows and lower low flows) as well as maintaining the seasonality of low- and high-flow periods (NRC, 2005).

After impoundment behind Kingsley Dam, the annual maximum discharge declined from about 450 m³/s to 175 m³/s, which has remained constant through 2019 (Figure 4.29). The historic flow regime of the Platte River annually had numerous no-flow days, with an average of 78 per year at Overton, NE (Schumm, Reference Schumm and Gupta2007). The local alluvial water table was too deep to support woody vegetation growth across the broad exposed dry channel bed surface. The inability of woody vegetation to colonize channel bars helped maintain a dynamic channel bed, with sedimentary bars frequently mobilized and reworked during higher flow periods (Johnson, Reference Johnson1994). Since impoundment there are zero no-flow days (at Overton, NE), as the flow regime has changed from intermittent to perennial (Schumm, Reference Schumm and Gupta2007).

Figure 4.29.

Historic changes in channel width and annual maximum peak discharge for the Platte River near Overton, Nebraska (USGS 06768000) in relation to upstream closure of Kingsley Dam in 1941 along North Platte River (~120 km downstream). Recent annual maximum discharge values are 104 (m³/s) and 297 (m³/s) for water years 2009 and 2019, respectively. Recent channel width value is 517 m for 2015 (September 27), measured between vegetated banks at 1 km intervals along a 5-km channel reach at Overton, NE in Google Earth Pro by author.

(Data source: Williams, 1978b, with author modifications.)

The seasonality of the high and low flow periods is also important, as prior to impoundment the timing of the premodified flood pulse coincided with the period of seed germination for Populus-Salix in June. Importantly, the high-flow disturbance period in late spring and early summer limited root growth and sedimentary bar colonization, as revealed by comparison of historic air photo with modern imagery (Figure 4.30). After impoundment and substantial water withdrawals for agriculture resulted in the lowering of seasonal flood pulses, Populus-Salix vegetation rapidly colonized formerly active channel bars and transformed the riparian environment (Johnson, Reference Johnson1994; NRC, 2005) (Figure 4.30).

Figure 4.30.

Comparison of historic and modern Platte River riparian landscape imagery (same area).

(A)

Historic (1938) vertical aerial photograph of the Platte River at Kearney. Numerous active sedimentary bars and few vegetated bars (islands). Distance along bridge (between vegetated banks) is 386 m.

(B)

Platte River at Kearney, NE September 27, 2015. Compare with 1938 photo. Distance along bridge between vegetated channel banks is 264 m.

(Sources: (A) U.S. Geological Survey vertical air photo, 9″ × 9″ (23 cm × 23 cm) at 1:10,000 scale, Platte River Program – Historical 1938 Aerial Photography 47_53, (B) similar area as A from Google Earth Pro.)

The scenario reviewed for the Platte River has played out across numerous riparian lands after upstream flow impoundment, particularly in dryland settings. In addition to changes to riparian sedimentary and vegetation environments, changes to fisheries are acute.

4.3.5.1 Yangtze Fishery Decline Downstream of Three Gorges Dam

Because of pollution and hydraulic engineering projects, the Yangtze River fishery was already in trouble prior to construction of Three Gorges Dam. Since closure of Three Gorges Dam in 2003, however, the fishery has continued to decline. The number of carp eggs and carp larvae has plummeted from about 3.5 billion in 1997 to less than 500 million by 2003 (Figure 4.31A). The situation has improved in the past decade with more environmentally friendly approaches to reservoir management, increasing to almost 1.5 billion carp larvae and eggs by 2015 (Cheng et al., Reference Cheng, Opperman, Tickner, Speed, Guo and Chen2018). An important management approach is to coordinate the reservoir release schedule with critical life cycle stages of key aquatic organisms. In the case of the Yangtze River it involves reintroducing lateral hydrologic connectivity between the river and the massive floodplain lakes, such as Dongting Lake, which serves as vital carp habitat (Yia et al., Reference Yia, Yang and Zhang2010; Ru and Liu, Reference Ru and Liu2013).

Figure 4.31.

Three different downstream responses of riverine fisheries to dam construction, including (A) decline in carp eggs and larvae on Yangtze River before and after closure (2003) of Three Gorges Dam, with subsequent environmental flows program after 2011; (B) Volga River decline in riverine and floodplain production after closure of Volgograd Dam in 1959; and (C) coastal fish catch before and after Aswan High Dam impoundment (1964) of the Nile in relation to increased fertilizer use for intensive agriculture.

(Sources: Volga River, Schmutz and Moog, 2018, with data from Górski et al., 2012; Yangtze, Cheng et al., 2018; Nile, Oczkowski et al., 2009.)

4.3.5.2 Downstream Impacts to Volga Fishery

Changes to the lower Volga River imposed by closure of Volgograd Dam (1959) reduced riparian spawning areas by ~80%, and is associated with large declines in fish populations, particularly the iconic Russian sturgeon (Acipenser gueldenstaedtii) (Secor et al., Reference Secor, Arefjev, Nikolaev and Sharov2000; Maltsev, Reference Maltsev, Carmona, Domezain, García-Gallego, Hernando, Rodríguez and Ruiz-Rejón2009). The annual fish catch in both the channel and the floodplain since impoundment of the lower Volga has plummeted (Figure 4.31B). In addition to woody vegetation encroachment, numerous side channels that served as prime aquatic habitat for fish spawning infilled with sediment (Middelkoop et al., Reference Middelkoop, Alabyan, Babich, Ivanov, Hudson and Middelkoop2015). In addition to impoundment, other factors related to the decline of the lower Volga fishery includes water pollution, water stress caused by extensive withdrawal for agriculture, and illegal fishing (Ruban et al., Reference Ruban, Khodorevskaya and Shatunovskii2019).

4.3.5.3 Nile Coastal Fishery and Industrial Agriculture

The impact of impoundment on downstream fisheries is quite different for the Nile River. Because of a 90% reduction in Nile flooding after closure of the Aswan High Dam (Oczkowski et al., Reference Oczkowski, Nixon, Granger, El-Sayed and McKinney2009), nutrients historically associated with the Nile flood pulse were impounded in Lake Nasser, resulting in a collapse of coastal primary productivity (Dorozynski, Reference Dorozynski1975). The loss of the “Nile bloom” resulted in an abrupt decline of the Nile delta fishery. The annual sardine catches, for example, declined from 37,000 tons (1962–1965) to 6,500 tons (1966–1970) (Nixon, Reference Nixon2003).

An unintended consequence of Egypt’s moves to industrialized agriculture and its booming delta urban population, however, caused the fishery to rebound in the 1980s. This occurred because of large increases in chemical fertilizers (phosphate and nitrogen) to support agriculture, which is consumed by a much larger urban coastal population. Annual fertilizer use increased almost fourfold, from about 340,000 tons to 1,300,000 tons (Figure 4.31C). The sequence is fairly straightforward. The fertilized agricultural products are consumed, and the raw sewage is discharged into the lagoons and marine environments, increasing primary productively. A substantial proportion (60–100%) of the Nile delta fishery is supported by nutrients from fertilizer that is largely supplied by sewage disposal (Oczkowski et al., Reference Oczkowski, Nixon, Granger, El-Sayed and McKinney2009). The Aswan High Dam is not the cause of the nutrification of the Nile delta waters, but it is part of a larger industrial agricultural system that utilizes large amounts of chemical fertilizers (Dorozynski, Reference Dorozynski1975). Unexpectedly, annual fish catch totals now far exceed historic levels (Nixon, Reference Nixon2003; Oczkowski et al., Reference Oczkowski, Nixon, Granger, El-Sayed and McKinney2009).

4.3.5.4 Mekong Fishery and Aquatic Megafauna

Fluvial riparian environments along the Mekong are being strangled by dams, while the delta is drowning because of sediment starvation (Syvitski et al., Reference Syvitski, Vörösmarty, Kettner and Green2005; Kummu and Varis, Reference Kummu and Varis2007; Kondolf et al., Reference Kondolf, Rubin and Minear2014b; Lu et al., Reference Lu, Kummu and Oeurng2014; Winemiller et al., Reference Winemiller, McIntyre and Castello2016; Best and Darby, Reference Best and Darby2020; Minderhoud et al., Reference Minderhoud, Middelkoop, Erkens and Stouthamer2020; Yoshida et al., Reference Yoshida, Lee and Trung2020). With 60 million living within 15 km of the lower Mekong, much of the population depends upon riparian services provided by a free-flowing river, including transportation, navigation, agriculture, commerce, and sustenance (Dugan et al., Reference Dugan, Barlow and Agostinho2010). The Mekong inland fishery is the world’s largest and produces some 2.0–2.6 million tons of fish per year, accounting for 49–82% of animal protein consumed by the Mekong population (Yoshida et al., Reference Yoshida, Lee and Trung2020). Proposed fish ladders and gates to facilitate fish migration around dams are inadequate to cope with the diversity of behavior exhibited by the numerous fish species (Dugan et al., Reference Dugan, Barlow and Agostinho2010; Ziv et al., Reference Ziv, Baran, Nam, Rodríguez-Iturbe and Levin2012; Yoshida et al., Reference Yoshida, Lee and Trung2020). As 40–70% of the Mekong fish are migratory, dam construction along the main-stem and tributaries blocks migratory fish, and threatens livelihoods. In addition to the main-stem channel, the large tributary channels in the middle and lower Mekong basin are also impounded.

A special aquatic megafauna that inhabits the lower Mekong is the Irrawaddy Dolphin (Orcaella brevirostris), which migrates between the estuarine delta and middle reaches of the Mekong. As with river dolphins in many impounded rivers, the Irrawaddy dolphin is heavily threatened (Winemiller et al., Reference Winemiller, McIntyre and Castello2016). Up to half of the remaining Mekong population of Irrawaddy Dolphin, estimated at only eighty individuals (Krützen et al., Reference Krützen, Beasley and Ackermann2018), were expected to be lost because of being in the vicinity of the construction process (e.g., earth works, explosions, dredging) of the Don Sahong hydroelectric dam (commissioned 2020), located in Laos 2 km from the Laotian–Cambodian border (Dugan et al., Reference Dugan, Barlow and Agostinho2010). A sliver of hope occurred in March 2020 when the Cambodian government abandoned large dams planned for the lower Mekong, and put a ten-year moratorium on any new dams on the Mekong. This is crucial because additional large main-stem dams in the lower Mekong would block upstream fish migration with the rich delta nurseries, as well as the Tonle Sap inland lake and wetland system, which in itself represents 60% of Cambodia’s annual fish catch (Johnstone and Sithirith, Reference Johnstone, Sithirith, Finlayson, Everard and Irvine2018).

4.3.6.1 Thermal Stratification

Thermal stratification in large reservoirs varies with changes in riverine inputs and temperature (Rashid, Reference Rashid, Crul and Roest1995; Jiang et al., Reference Jiang, Wang and Ni2018). Dissolved oxygen levels in Lake Nasser and Nubia, for example, vary seasonally and spatially along the ~500-km long reservoir. Reservoir stratification is pronounced in July and August with dissolved oxygen levels averaging between 9.5 mg/L at the surface and 0.0 mg/L at a depth of 15 m. In the late autumn and winter, the water is mixed and does not exhibit stratification and the reservoir is oxygen saturated from the surface to bottom (Rashid, Reference Rashid, Crul and Roest1995). Large reservoirs in the lower Mekong basin have higher temperature layers (Figure 4.33) that commonly range from 28°C to 30°C and somewhat “float” upon denser cooler layers (Hortle and Nam, Reference Hortle and Nam2017). In cooler climatic settings, head-of-reservoir spring and summer water temperatures can be too warm and detrimental to the life cycle of aquatic organisms, particularly concerning spawning habitat for fish accustomed to cooler water temperatures.

Figure 4.33.

Reservoir depth (m) profile for water temperature (C) and dissolved oxygen (mg/L) in 2004 for stratified Nam Leuk Reservoir, Laos, Mekong basin.

(Source: Hortle and Nam, 2017.)

For decades these issues were documented for a range of reservoir types across an international range of environments (Young et al., Reference Young, Kent and Whiteside1976; Rashid, Reference Rashid, Crul and Roest1995; Kunz et al., Reference Kunz, Wüest, Wehrli, Landert and Senn2011), and are of pressing concern considering the current boom in large dam construction (Winton et al., Reference Winton, Calamita and Wehrli2019).

4.3.6.2 Reservoirs and Phosphorus Storage

Because many dams are built to support irrigation and agriculture it is not surprising that increases in agriculture over the twentieth century resulted in increased nutrient loading of reservoirs. Globally, 12% of total phosphorus loads transported by rivers were trapped in reservoirs in 2000, a 5% increase over 1970 levels (Table 4.10). While total river phosphorus loads modestly increased between 1970 and 2000, the amount of phosphorus retained in reservoirs effectively doubled, as total phosphorus retained in reservoirs increased from 22 to 42 Gmol/yr between 1970 and 2000. Thus, the much higher amount of retention is due to the increase in reservoir storage capacity that occurred over this period (e.g., Figure 4.4). And the ongoing massive dam construction boom that will result in hundreds of new large reservoirs is projected to result in further increases in phosphorus storage, with some scenarios projecting an increase from 12% to 17% by 2030 (Maavara et al., Reference Maavara, Parsons and Ridenour2015). It should be noted that riverine phosphorus budgets (Table 4.10) also reveals an alarming increase in phosphorus bypassing reservoirs, which contributes to coastal eutrophication and the proliferation of hypoxic dead zones at the mouths of rivers around the world (Breitburg et al., Reference Breitburg, Levin and Oschlies2018).

Table 4.10Global retentions of total phosphorus and reactive phosphorus* by reservoirs for 1970, 2000, and 2030

^* Reactive phosphorus defined as sum of total dissolved phosphorus, exchangeable phosphorus, and particulate organic phosphorus and is considered the fraction of total phosphorus potentially bioavailable.

^** GO = Global Orchestration scenario. See Alcamo et al. (2006) for description of millennium ecosystem assessment scenarios.

Source: Maavara et al. (Reference Maavara, Parsons and Ridenour2015).

The Mississippi basin has been the global leader in reactive phosphorus retained in reservoirs since 1970, with about 50% of its phosphorus load retained in its 700 odd reservoirs (Table 4.11). The amount of reactive phosphorus retained in reservoirs in 2000 was 920 × 10⁶ mol/yr and by 2030 will undergo a modest increase. But, by 2030 the Mississippi will be replaced by the Yangtze River as the highest reservoir retention of reactive phosphorus, which at 2,898 × 10⁶ mol/yr will represent only 35% of its total load (Table 4.11). Other rivers projected to enter into the top 10 rankings include the Paraná, Huanghe, Zaire, and Mekong Rivers. These rivers are not only undergoing increased dam construction, but also are located in regions undergoing increased land cover change for mechanized agriculture that is reliant upon artificial fertilizers, specifically high amounts of phosphorus. The Yangtze River, for example, is projected to see an enormous increase in reactive phosphorus from 3,758 × 10⁶ mol/yr in 2000 to 8,327 × 10⁶ mol/yr by 2030. Interestingly, the Zambezi River basin will remain among the top ten for phosphorus retention, although it has by far the lowest phosphorus loads. This is because the high trap efficiency of its reservoirs, increasing from 62% to 73% between 2000 and 2030, which includes Kariba Reservoir, the world’s largest reservoir by volume (Kunz et al., Reference Kunz, Wüest, Wehrli, Landert and Senn2011; Giosan et al., Reference Giosan, Syvitski, Constantinescu and Day2014; Maavara et al., Reference Maavara, Parsons and Ridenour2015; Winton et al., Reference Winton, Calamita and Wehrli2019).

Table 4.11Ranking of basins by reservoir storage of reactive phosphorus for 2000 and a projected scenario for 2030

Source: Maavara et al. (Reference Maavara, Parsons and Ridenour2015).

4.3.6.3 Downstream Changes to Temperature and Oxygen

Dam and reservoir management strategies influence water quality within reservoirs and downstream of dams. In this regard, a fundamental consideration is the depth of the reservoir thermocline/oxycline relative to the spillway intake. Many large reservoirs have outflow levels deeper than ~10 m, which is usually below the thermocline, even during warm and dry conditions (Young et al., Reference Young, Kent and Whiteside1976; Winton et al., Reference Winton, Calamita and Wehrli2019). The release of potentially cold, hypoxic, and nutrient-enriched waters can have detrimental consequences to downstream riparian environments. Cooler dam outflow water along the Yangtze River resulted in a delay of fish spawning by several weeks (Zhong and Power, Reference Zhong and Power2015). Macroinvertebrate communities along Guadalupe River exhibited qualities consistent with ecological stress, with reduced diversity and higher numbers of fewer species downstream of Canyon Dam, a deep storage reservoir in the Texas Hill Country (Young et al., Reference Young, Kent and Whiteside1976).

Outflow from large reservoirs is clear and cold with lower oxygen levels. Because of the depth and size of reservoirs, the volume of outflow potentially alters downstream aquatic environments for tens of kilometers. This is shown for two of the larger tributaries to the lower Mekong (i.e., Figure 4.19). Streamflow temperatures varied by four and six degrees (Celcius) 56 km and 78 km downstream of recently constructed dams along the Sesan and Srepok Rivers, respectively (Figure 4.34). The streamflow for both rivers underwent substantial cooling during the dry season, which persisted into the first half of the wet season (May, June, July). Thus, reduced streamflow variability caused by dams occurs with temperature, as it does with water level (stage). The importance of this is highly seasonal and related to the life cycles of aquatic organisms, including fish behavior related to feeding and spawning (Bonnema et al., Reference Bonnema, Hossain, Nijssen and Holtgrieve2020).

Figure 4.34.

Downstream impact of lower Mekong basin dams on water temperature. Comparison of pre-dam (2004–2008) and post-dam (2009–2011) average monthly streamflow temperatures for the (A) Sesan and (B) Srepok Rivers, with dam construction in 2008 and 2009. No major dams were constructed on the (C) Sekong River over the study period, which provides a “control” for comparison. Temperature monitoring stations along the Sesan and Srepok Rivers were 56 km and 78 km downstream of dams, respectively. See Figure 4.19 for location of rivers within the lower Mekong basin.

(Source: Bonnema et al., 2020.)

4.3.6.4 Reservoir Water Quality Management: Tennessee Valley Authority Case Study

The Tennessee Valley Authority (TVA) is a massive federal economic development project in the southeast United States that was initiated during the Great Depression of the 1930s (Figure 4.35). The main focus of the TVA was to build dams for hydroelectricity, flood control, and navigation. The TVA operates forty-nine dams, with twenty-nine built for hydroelectricity and twenty non-power-generating dams with a primary purpose of flood control. The dams are located within the Tennessee River basin (105,900 km²), which drains the southwestern Appalachian Mountains and flows westerly, joining the Ohio River ~40 km upstream of its confluence with the Middle Mississippi to form the lower Mississippi. The Tennessee River is heavily impounded and includes nine main-stem dams and locks that create a 1,050-km long navigable corridor. The approach to reservoir management upon completion of the dams was to maximize hydropower, which resulted in storage and release schedules that did not coincide with riparian ecosystems; essentially the opposite of environmental flows.

Figure 4.35.

Map of the profile of the Tennessee River, showing locations of TVA dams and drainage network of the Tennessee River basin. Unit conversion: 1 mile = 1.6094 km, 1 ft = 0.3048 m.

(Source: Tennessee State Library and Archives.)

Reservoir outflow for many dams built by the TVA had poor water quality that did not meet established environmental standards (Table 4.12). A study of twenty dams found that over a thirty-six-year period, downstream flow volumes were too low to support critical riparian functions of aquatic habitat (Higgins and Brock, Reference Higgins and Brock1999). The downstream distance where streamflow volumes were too low exceeded 20 km for ten dams. The Holston River downstream of Cherokee Dam has an average discharge of 130 m³/s, but for a distance of 76 km below, the dam only averaged a minimum flow of 2 m³/s. This effectively resulted in a dry channel bed, although the region has among the highest annual rainfall totals in the continental United States. Additionally, sixteen of the TVA dams had outflow with dissolved oxygen levels below environmental standards (Table 4.12). The French Broad River, downstream of Douglas Dam, had average minimum dissolved oxygen levels of just 0.9 mg/L for an average of 113 days per year over a distance of 129 km downstream of the dam (Higgins and Brock, Reference Higgins and Brock1999). Such low levels of dissolved oxygen represent hypoxic conditions that can be fatal to immobile benthic macroinvertebrates while fish abandon the segment. The river effectively becomes a riparian dead zone.

Table 4.12Reservoir release flows and dissolved oxygen (DO) levels for selected reservoirs in the Tennessee Valley Authority project

Note: Based on daily streamflow (m³/s) and weekly DO (mg/L) data from 1960 through 1996. Data affected by release improvements were omitted.

^* DO target is 6.0 mg/L for nine dams with cold water downstream fishery, and 4.0 mg/L for eleven dams with warmwater downstream fishery.

^** Tailwaters not impacted by low flows due to backwater from downstream reservoirs.

Source: Higgins and Brock (Reference Higgins and Brock1999).

Fortunately, in 1991 the TVA initiated changes to its reservoir management strategy to improve downstream ecological conditions (Bednarek and Hart, Reference Bednarek and Hart2005). To increase the wetted surface area of the channel bed, measures included turbine pulsing to create a steady outflow, increased minimum flow volumes, and construction of low weirs downstream to maintain water levels at critical heights. To increase dissolved oxygen levels, modifications were made that involved injecting oxygen into reservoir water as well as structures immediately below the dam to create turbulence and aeration of reservoir outflow water. The measures were largely deemed effective, and both abiotic and biotic conditions improved downstream of the dams. Average dissolved oxygen levels increased by ~34%. The average increase in minimum velocity and discharge was 59% and 528%, respectively. While it was difficult to discern the importance of increased flow velocity relative to increased dissolved oxygen, macroinvertebrate assemblages significantly improved across the study segments. The number of macroinvertebrate families increased by 36% and there was a 13% reduction in the number of macroinvertebrates tolerant of low water quality conditions (Bednarek and Hart, Reference Bednarek and Hart2005).

Downstream rehabilitation of impounded rivers by modifying dam and reservoir operations is critically important because of increased stressors of land cover change and climate change, and because many dams will effectively remain a permanent component of the riparian landscape. But designing reservoir outflow for environmental flows remains a work in progress (Bednarek and Hart, Reference Bednarek and Hart2005; Rader et al., Reference Rader, Voelz and Ward2008; Kunz et al., Reference Kunz, Senn, Wehrli, Mwelwa and Wüest2013). While the efforts of the TVA to improve the downstream environmental integrity of riparian environments were deemed somewhat successful, the TVA’s highly regulated rivers far from resemble a natural stream, and restoration was not the goal. Riparian ecologists have developed an intricate knowledge of critical life cycle phases of stream biota and their relation to abiotic components (e.g., bed material, flow velocity, depth, temperature). Such knowledge has resulted in conceptual models that provide useful signposts to guide river managers in environmentally effective dam and reservoir management, including lateral connectivity with floodplains (Ward and Stanford, Reference Ward and Stanford1995). The actual success of such mitigation measures, however, varies greatly. In view of each river and reservoir representing a unique combination of natural and human controls, such management measures are only likely to be effective with a detailed knowledge of the local fluvial environment (Robinson et al., Reference Robinson, Uelinger and Monaghan2003).

4.4.1.1 Legislation and Policy Instruments

Government institutions across international, federal, state/provincial, and local scales have varying types of policy instruments that pertain to specific drivers of dam removal. The U.S. Clean Water Act (Sec. 404) and Endangered Species Acts, for example, are federal legislative tools that can be used in environmentally motivated dam removal ‘projects’, although they are seldom employed (Bowman, Reference Bowman2002; Doyle et al., Reference Doyle, Stanley and Harbor2003b; Opperman et al., Reference Opperman, Royte, Banks, Day and Apse2011). The relicensing of a dam requires a review of the Endangered Species Act as it pertains to adverse environmental impacts as well as measures to manage and mitigate adverse impacts to riparian environments. Sweden has the federal Environmental Objective, which identified sixteen ecosystem goals to be attained by 2020. This included a stated priority of maintaining healthy rivers, lakes, groundwater, and wetlands, among others. A specific facet of the objective was to restore 25% of valuable rivers by 2010, a stimulus for dam removal efforts in Sweden based on environmental justifications (Lejon et al., Reference Lejon, Renöfält and Nilsson2009). For rivers that intersect international borders it is important to have international agreements and institutions to specify cooperation in support of healthy rivers.

The European Union’s Water Framework Directive (WFD) requires that all EU nations manage water in the context of sustainability. Indeed, Article 4 of the WFD includes specific ecologic and hydrologic targets that nations are supposed to attain by 2027 (Directive 2000/60/EC). To adapt to the WFD, the EU developed the River Basin Management Plan as a common strategy with specific deadlines to meet each of the WFD requirements. While the requirements were in accord with the concept of dam removal to improve riparian health, a number of nations did not view the WFD in these terms and were slower to pursue dam removal projects. The implementation and enforcement of the EU’s WFD in the context of dam removal was recently strengthened (May 2020) by the EU Biodiversity Strategy for 2030. The Biodiversity Strategy requires restoration of at least 25,000 linear km of rivers to a free-flowing state through the removal of barriers (dams and weirs), in addition to floodplain and wetland restoration.

4.4.1.2 Economics and Dam Removal

Economics is a vital driver of dam removal and includes multiple facets. While revenue associated with electricity generation is often an important primary consideration for hydropower dams, other economic considerations include revenue associated with tourism and recreation, land use and associated income generation, property values, taxes, fisheries production, sediment management, and dam maintenance and repair (Lawson, Reference Lawson2016). Dam removal projects include formal cost–benefit analyses, which compare the cost of removal to the economic benefits of removal as well as the cost of alternatives to dam removal. Alternatives to dam removal often include the cost of upgrading to meet new standards as regards safety, mechanical and electrical generation, and environmental mitigation, particularly fish passages. The alternatives to dam removal are often twofold higher than the cost of dam removal (Lawson, Reference Lawson2016). Sediment management (discussed below) is usually the most expensive part of dam removal. For some dams, changes to residential property values can be a larger consideration. In some instances, property values increased after dam removal for rivers in Maine and Wisconsin (United States), although it depends on the proximity to the river (Lewis et al., Reference Lewis, Bohlen and Wilson2008; Provencher et al., Reference Provencher, Sarakinos and Meyer2008; Auerbach et al., Reference Auerbach, Deisenroth, McShane, McCluney and Poff2014; Lawson, Reference Lawson2016). And each economic factor may be used to make a case for or against dam removal depending on the competing stakeholders (Morris and Fan, Reference Morris and Fan1998; WCD, 2000; Heinz Center, 2002).

After a decades-long dispute about removing four dams along a 225 km stretch of the lower Snake River, Washington (Ice Harbor, Lower Monumental, Little Goose, and Lower Granite dams) that provide power, flood control, water for irrigation, and navigation, but only about 5% of the regional electricity, the U.S. EPA in 2020 decided not to remove the four dams. The stated reason was concern that a loss of power generation would increase electricity rates for consumers. Earlier economic analyses had shown a net benefit of $310 million, and that the river recreation gains actually exceeded lost revenue from reservoir recreation (Loomis, Reference Loomis2002). The dams ensure ~450 km of navigation from the Pacific Ocean at the mouth of the Columbia River to Lewiston, Idaho. But the dams are also associated with steep decline in Pacific salmon stock as well as declines in species that heavily rely upon salmon for feeding, which includes Southern Resident killer whales.

Globally, numerous dams are older than five decades (e.g., Figure 4.4) and in need of costly repairs, maintenance, and upgrading (Morris and Fan, Reference Morris and Fan1998; Heinz Center, 2002; Doyle et al., Reference Doyle, Stanley and Havlick2008). The average age of US dams is fifty-seven years (as of 2020), and over 10,000 dams (>2 m) are older than seventy years (e.g., Figure 4.8). The cost of upgrading and repairing such infrastructure is enormous, and was estimated in 2016 at about $65 billion (Table 4.14). Because there are so many old dams with questionable structural integrity, an additional strong motivation for dam removal is liability, which ultimately relates to economics (Heinz Center, 2002). Most dams are locally owned and many owners are increasingly opting for dam removal to avoid expenses associated with liability or repairs. Small dam removal projects often cost less than the cost of safety repairs (Born et al., Reference Born, Genskow, Filbert, Hernandez-Mora, Keefer and White1998; Bowman, Reference Bowman2002). Thus, the same rationale for the original construction of dams – economics – is now being utilized to justify their removal.

Table 4.14Cost to rehabilitate US dams

Source: ASDSO (2016).

4.4.3.1 Dam Removal in North America

Dam removal is occurring at an accelerated pace and being guided by increasingly sophisticated procedures and monitoring conventions, although there is great regional and national and international variability across the United States, Canada, and Mexico.

In total, 1,699 dams were removed in the United States between 1900 and 2019 (Figure 4.37). And 806 dams were removed over the past decade, nearly double the previous decade. Most dams removed are less than 5 m in height. In the United States, the hotspots for dam removal are in the northeast, northwest, and the northern Midwest (American Rivers, 2020). Numerous mill dams in New England have been removed. Many of these are well over a hundred years old and represent the region of the most intense and oldest river impoundment in the United States (Walter and Merritts, Reference Walter and Merritts2008). The largest dam removal project in New England is the Great Works Dam in 2012 and the Veazie Dam in 2013 along the lower Penobscot River (22,196 km²) in Maine. The project also included modification to dam structures to create fish passages, and was driven to restore some 3,200 km of riparian habitat to anadromous fish, such as Atlantic salmon (Salmo salar), alewife (Alosa pseudoharengus), American shad (Alosa sapidissima), and Atlantic sturgeon (Acipenser oxyrinchus oxyrinchus). And it worked, as fish populations quickly returned within a few years, with anadromous fish species increasing while slower lacustrine species declined (Watson et al., Reference Watson, Coghlan, Zydlewski, Hayes and Kiraly2018). California had twenty-three dams removed in 2019, and Pennsylvania had fourteen. Texas had one dam removed in 2019, and has only had eight in total removed. The paucity of dam removal projects in Texas is concerning because it has the highest number of dams and the largest proportion of its surface water impounded of any state in the United States (Graf, Reference Graf1999).

Figure 4.37.

US dam removal between 1900 and 2019. Total dams removed are 1,699 (May 22, 2020). Dam removal between 1900 and 1979 grouped in twenty-year periods and from 1980 to 2019 by decade.

(Author figure. Data source: American Rivers, 2020.)

In terms of height, 349 dams that are higher than 5 m have been removed in the United States. In the Pacific Northwest, dams are being removed in smaller coastal draining rivers where it is mainly about restoration of Pacific salmon habitat (all five species), which quickly return to spawn in upstream rivers for the first time after many decades (Lierman et al., Reference Ma, Nittrouer and Naito2017). The largest dam removal project in the world – to date – occurred along the Elwha River (820 km²) in the Olympic Peninsula of coastal Washington over a couple of years (see discussion below). The removal project occurred in two phases, the 33-m Elwha River Dam in 2011 followed in 2012 by a phased removal of the 64-m Glines Canyon Dam (Ritchie et al., Reference Ritchie, Warrick and East2018).

Dams are also being removed in Canada, although at a slower pace than the United States. The removal of small dams in Canada is tempered by the fact that large dams have recently been completed and are still being constructed along major waterways. Since 2006 twenty-one dams have been removed in British Columbia, including four large dams >9 m in height. Dam removal is proceeding within established removal procedures and guidelines. A positive sign that dam removal is becoming a mature environmental industry in Canada is that sophisticated guidelines are being developed within clear policy frameworks at the scale of individual provinces. Ontario, for example, has the Lakes and Rivers Improvement Act (LRIA) that specifically includes a range of fluvial geomorphic criteria to develop a dam removal plan, including sedimentary and morphologic adjustments (Ontario Ministry of Natural Resources, Reference Dettinger2011).

In Mexico, dam removal is not occurring with any documented frequency. Indeed, the trend is somewhat the opposite and several large dams were constructed in the 2000s. In 2020, Mexico confirmed it will go forward with the massive Chicoasén II hydroelectric dam (240 MW) in the deep canyons of the Grijalva River in the mountainous southern state of Chiapas. The dam is the second phase of a larger hydroelectric site that includes the 261-m high Manuel Moreno Torres (Chicoasén) Dam, the tallest dam in North America. Several other proposed Mexican dams, fortunately, have been stalled or canceled over the past decade because of substantial opposition from environmentalists, human rights supporters, and indigenous peoples.

4.4.3.2 Dam Removal in the European Union

Dam removal in the European UnionFootnote ³ is occurring with much greater frequency over the past decade, with France leading the way (Fernández Garrido, Reference Fernández Garrido2018; Schiermeier, 2018). The increasing pace of dam removal is occurring for several key reasons, including: (1) The European Union Water Framework Directive (WFD), which requires free-flowing rivers to attain good ecological status by 2027 (noted above); (2) environmentally oriented national strategies supported with science that view river networks as linked to landscapes; and (3) the development of effective nongovernmental organizations that champion the cause of dam removal. Nevertheless, like the United States there is considerable regional variability in the pace of dam removal. Many of the debates and the roles of stakeholders with regard to dam removal projects are similar to the United States (Born et al., Reference Born, Genskow, Filbert, Hernandez-Mora, Keefer and White1998; Lejon et al., Reference Lejon, Renöfält and Nilsson2009; Jorgensen and Renofalt, Reference Jorgensen and Renofalt2013; Magilligan et al., Reference Ma, Nittrouer and Naito2017). To obtain a clearer picture of dam removal across the EU it is crucial that a standard system of data reporting is developed.

A recent “fitness check” of the effectiveness of implementation of the EU Water Framework Directive revealed that the quality of some 40% of European rivers is under “hydromorphological” stress caused by dams and hydraulic engineering (EEA, 2018; EC, 2019). This recognition further increased attention to the issue of dam removal, as well as passage in 2020 of the EU Biodiversity Strategy for 2030 that includes a stated purpose to greatly increase the length of free-flowing rivers in Europe. Dam Removal Europe (damremoval.edu) is the key NGO, and has singularly focused on removing barriers (dams and small weirs) that are obstacles to fish migration, which like North America, is the primary driver of dam removal. A recent development (as of 2018) is the formation of AMBER (Adaptive Management of Barriers in European Rivers), a consortium of twenty key private and public specialists that represent a range of stakeholders involved with dam removal in Europe. AMBER includes partners from the hydropower industry, river authorities, nongovernmental organizations, and university scientists with the intention to enhance collaboration and knowledge transfer about the process of dam removal. An important task being undertaken by AMBER is to arrive at a tally and site-specific characterization of dam removal projects, which is difficult because of different protocols for reporting, as noted for Spain. Additionally, an important element of AMBER is “citizen science,” which directly involves the public in data collection and dissemination. Combined, these two thrusts of European dam removal have increased the visibility of the topic to a range of public and private stakeholders, increasing public participation in the process of dam removal.

According to the DamRemoval.eu NGO, some 5,000 dams (and weirs) have been removed across five European nations as of 2018. This includes France with 2,425 dam removals (partially removed obstacles: 5,728), Sweden with >1,600, Finland with >450, Spain with >250, and Great Britain with 156 dam removals (Fernández Garrido, Reference Fernández Garrido2018). Many EU nations have ambitious plans for dam removal that link national strategies with the EU WFD. Among the several significant dams removed in Europe are the Retuerta dam (14 m) and the La Gotera dam (8 m) in the Duoro basin (Spain) in 2013 and 2011, respectively. The latter was initiated as part of the Spanish National Strategy of River Restoration to increase hydrologic continuity through the Alto Bernesga Biosphere Reserve of the Man and Biosphere Programme. In France’s nationally symbolic Loire basin, the Saint Etienne du Vigan dam (14 m) on the Allier tributary and the Maisons Rouges dam (4 m) on the Vienne tributary were both removed in 1998 as part of the French governments “Plan Loire Grandeur Nature.” The main focus is to increase fish migration to the upper basin headwaters, particularly for spawning of Atlantic salmon (Salmo salar). The largest dam removal project in Europe occurred along the Sélune River in Normandy, France. The 36-m high Vezins Dam was the first of two dams to be removed (2019), with the second being the 15-m high La Roche Qui Boit removed in 2021. Combined, the removal of the two dams reopen 91 km of a coastal draining river to migrating Atlantic salmon, eels, and other riparian wildlife, which empties into the bay of Mont-Saint-Michel – a UNESCO world heritage site and key stakeholder related to cultural heritage.

Although progress is being championed in scientific media (Schiermeier, Reference Schiermeier2018), dam removal varies greatly across the European Union. And it often varies within individual nations between states/provinces or basin, as with the issue of data reporting in Spain. Data reporting procedures in Spain are not centrally coordinated and instead are organized at the scale of the larger drainage basins, which vary considerably in format and in public accessibility (Rincón Sanz and Gortázar Rubial, Reference Rincón Sanz and Gortázar Rubial2016). In the Douro River basin (98,400 km²), for example, 116 dams and barriers were removed between 2009 and 2016, the highest in any basin in Spain. Of these dams, 31 were <2 m height, 16 were 2–5 m height, 2 were >10 m height, but 67 of the dams removed in the Duoro basin did not include a recorded dam height or descriptive information (Ministry for Ecological Transition, MITECO, 2020). In comparison to the Duoro Basin, which has 145 large dams, the Ebro River basin (80,093 km²) is the most impounded in Spain, and has 299 large dams. But data from the Ebro River basin authority only reported 5 dams removed (Rincón Sanz and Gortázar Rubial, Reference Rincón Sanz and Gortázar Rubial2016).

4.4.4.1 Managing Reservoir Sediment Removal

Reservoir drawdown is a crucial step in dam removal, and management of sedimentary deposits during dam removal follows a sequence of specific steps (Figure 4.38). The method of drawdown is of key importance to sediment mobilization during and after reservoir drawdown. Rapid reservoir drawdown results in considerable hydraulic flushing and scour of reservoir deposits (Scheueklein, Reference Scheueklein1990), and produces a large sediment pulse that can be detrimental to downstream riparian environments (Grimardias et al., Reference Grimardias, Guillard and Cattanéo2017). A slower drawdown is more benign to downstream riparian environments, although it can result in a larger volume of reservoir deposits that are disconnected from active fluvial processes. Additionally, upon drawdown a sharp knickpoint forms near the dam base (Neave et al., Reference Neave, Rayburg and Swan2009). Subsequent upstream headcut migration of the knickpoint through reservoir deposits further removes sediment by triggering mass wasting and erosion of noncohesive deposits (Morris and Fan, Reference Morris and Fan1998; USSD, 2015).

Because reservoir deposits sequester large amounts of phosphorus and nitrogen from upstream agricultural landscapes (Maavara et al., Reference Maavara, Parsons and Ridenour2015), the release of nutrients into the active fluvial system and eventually into downstream coastal waters is a concern during the dewatering phase of dam removal (Riggsbee et al., Reference Riggsbee, Wetzel and Doyle2012).

Many dams impound polluted reservoir deposits and it is important that dam removal does not result in contaminated sediment pulses impacting downstream environments (Evans, Reference Evans2015). Sediment pollution from dam removal is of particular concern in reservoirs downstream of old industrial and mining areas (Tullos et al., Reference Tullos, Collins and Bellmore2016). In such cases, removal of contaminated deposits must precede dam removal, or dam removal may be untenable. Along the Hudson River in New York, the 1973 removal of Fort Edward Dam resulted in several hundred thousand cubic meters of sediment contaminated with polychlorinated biphenyls (PCBs) to be transported downstream. After decades of legal disputes with industry about whether the polluted sediment should be capped or removed, some 2 million m³ of sediment was eventually dredged from the river bed, in addition to floodplain restoration, at a cost of $561 million (Evans, Reference Evans2015). And legal disputes between the state of New York and industry are ongoing as to whether the polluted sediment has been satisfactory removed.

4.4.4.2 Removal of Toxic Reservoir Deposits

The Milltown Dam removal project along Clark Fork (59,320 km²), the largest river in Montana, included reservoir deposits polluted with arsenic, copper, and other metals associated with a legacy of mining (Sando and Vecchia, Reference Sando and Vecchia2016). The dam had been constructed in 1907 and within months an enormous flood (~500-year RI) deposited tons of polluted sediment into the reservoir (Evans, Reference Evans2015). The reservoir is part of Clark Fork Basin Superfund Complex, the largest superfund site in the U.S. (U.S. EPA, 2016). The environmental and health problems associated with the contaminated deposits had been known since the early 1980s when residents started to become ill from polluted drinking water. The issue came to a head in the winter of 1996 when a large ice jam threatened to breach the dam. To prevent Milltown dam from being damaged by the ice jam the reservoir water level was lowered. But the reservoir drawdown was too abrupt and resulted in hydraulic scour of polluted reservoir deposits, releasing a toxic pulse that resulted in a massive fish kill downstream (Diamond, Reference Diamond2006; Evans, Reference Evans2015). Prior to finally dismantling the 12.8 m tall dam in spring 2008 (initial breach March 28), some 2.2 million m³ of polluted deposits (40% of total reservoir volume) were hauled away by train and buried outside of the riparian zone (U.S. EPA, 2016). By far the largest proportion of the Milltown Dam removal cost was managing reservoir sediment removal ($120 million).

The environmental disaster associated with the abrupt 1996 drawdown of the Milltown dam on Clark Fork River provided a useful lesson employed in the dam removal strategy. Rather than an abrupt drawdown, reservoir drawdown occurred very slowly over nearly two years, extending from June 1, 2006 to the dam breach, on March 28, 2008. While arsenic levels increased during the drawdown, it increased much less than copper (Figure 4.39). Nevertheless, despite the attempt to limit pollutant mobilization during the drawdown, the increased suspended sediment concentrations during the drawdown period (June 2006–March 2008) flushed 140,000 metric tons of particulates, which included 44 tons of copper and 6.4 tons of arsenic eroded from the reservoir. This is interesting because it shows that even when exercising great caution with reservoir drawdown, it should reasonably be expected that fine-grained deposits will be hydraulically excavated from the reservoir. Additionally, higher sediment loads during the drawdown phase were also attributed to structural landscape modifications (roads, facilities, excavation, etc.) associated with preparation of dam removal. Following the dam breach, an additional 420,000 tons of suspended sediment were transported downstream, which included 15.8 tons of arsenic and 169 tons of copper. A discharge of 250 m³/s during the removal phase resulted in a sediment load of 20,000 while several months later the same discharge produced a sediment load of about 2,000 tons (Sando and Lambing, Reference Sando and Lambing2011). Overall, copper, arsenic, and suspended sediment decreased by 53%, 29%, and 22%, respectively, between the start of water year 1996 and the end of water year 2015 (Sando and Vecchia, Reference Sando and Vecchia2016).

Figure 4.39.

Arsenic, copper, and suspended sediment between 1996 and 2015 for Clark Fork River. Sampling was 4.8 km downstream of Milltown Dam. Note: y-axis is logarithmic (USGS station i.d.: 12340500).

(Source: Sando and Vecchia, 2016.)

4.4.4.3 Reworking of Reservoir Deposits

To evaluate changes to fluvial sediment budgets associated with dam removal it is necessary to calculate downstream sediment loads associated with the reworking of reservoir deposits. Sediment mass (tons) is commonly estimated from the volume (m³) of reservoir deposits (Randle and Bountry, Reference Randle and Bountry2017). Accurate sediment load estimates from reservoir deposits require data on the weight of specific particle size classes, and whether they are subaqueous (submerged) or subaerial (above the water surface) deposits. This data is obtained from reservoir sedimentation surveys that measure the bathymetry and topography of reservoir deposits, as well as the particle size distribution obtained from borings and geophysical mapping (Yang, Reference Yang2006).Footnote ⁴ The weight of fine-grained deposits varies considerably according to whether it is submerged or above water (Table 4.15). Additionally, cohesive deposits, such as clay, substantially compact over time, which reduces pore space and increases sample weight. For example, after fifty years of compaction clay loses 35% of its original volume (Figure 4.40), which also increases reservoir storage capacity.

Table 4.15Comparison of submerged and subaerial weight of reservoir sediment

Source: Randle and Bountry (Reference Randle and Bountry2017), based on Morris and Fan (Reference Morris and Fan1998).

Figure 4.40.

Reservoir sediment compaction over time for different particle size classes over one hundred years. Values are for continuously submerged deposits. Sand not shown because of minimal consolidation.

(Source: Annandale et al., 2016.)

Most reservoir deposits reassimilated into the active fluvial system are eroded within the first year or two of dam removal, particularly for smaller elongated reservoirs with coarse deposits (Major et al., Reference Ma, Nittrouer and Naito2017). After the first couple of years, the amount of reservoir sediment reworking abruptly decreases as the channel stabilizes and as reservoir deposits become anchored with vegetation. A review of twenty-one dam removal projects in the United States (Table 4.16) revealed that over 40% of reservoir deposits were reworked and resupplied to the active fluvial system within the first year of dam removal by moderated streamflow events (Major et al., Reference Ma, Nittrouer and Naito2017). The proportion of reservoir deposits that eventually become assimilated back into the active sediment regime varies according to several factors, including sediment texture (fine or coarse), lateral channel dynamics, flood regime, and the size of the reservoir relative to the river channel. The amount of reservoir reworking appears to be less for broader reservoirs with finer-grained deposits and less lateral channel activity (Doyle et al., Reference Doyle, Harbor and Stanley2003a), although more documentation of dam removal projects is needed across lowland settings with fine-grained reservoir deposits.

Table 4.16Reservoir sediment eroded after dam removal for US case studies

^* Approximately 40% of sediment excavated prior to dam breaching was part of Superfund remediation efforts.

^** Dam height from Sacklin and Ozaki (Reference Sacklin and Ozaki1988).

Data: Randle and Bountry (Reference Randle and Bountry2017).

An additional factor that influences reservoir sediment reworking includes the biogeographic regime relative to the seasonality of reservoir drawdown (Cannatelli and Crowe-Curran, Reference Cannatelli and Crowe-Curran2012). If vegetation quickly colonizes reservoir deposits following reservoir drawdown, sedimentary deposits can effectively remain “locked in place” within the old reservoir site. This can be a desirable outcome. If not, vegetation may have to be removed to increase the possibility of sediment reworking (Bountry and Randle et al., Reference Randle and Bountry2017). Channel incision within the reservoir deposits forms anthropogenic terraces, reducing the possibility of erosion and reworking during flood events. Additionally, some reservoir sediments deposited atop natural terraces are unlikely to be hydrologically connected during subsequent flood events. In some cases, temporary training dikes can steer laterally active channels toward reservoir deposits to increase their potential for erosion. Fine-grained reservoir deposits distributed across wide valleys are less likely to be reassimilated into the active fluvial system.

4.4.5.1 Elwha River Dam Removal Case Study

The Elwha River dam removal project in the US Pacific Northwest is an ideal “laboratory” to study the impacts of dam removal on geomorphic and related ecosystem processes for a coarse-grained system. This is because it is the world’s tallest dam removal project and occurred on a modest-sized (833 km², Q_-avg. 42.7 m³/s) mountainous (peak 1,452 m elev.) gravel bed river that drains directly to the coast. And Elwha River and Glines Canyon Dams were only 7.4 and 21.6 km upstream of the river mouth at the Juan de Fuca Strait, respectively. Thus, detectable geomorphic and ecological changes to downstream riparian and coastal environments was all but assured. The study includes a range of morphologic, biological, and hydraulic, and sedimentologic sampling within a conventional research framework to capture the disturbance (spatially: upstream, at reservoir, downstream; temporally: before, during, and after).

The Elwha River Dam removal project is an example of stakeholders joining forces to dismantle dams for river restoration. The S’Klallam people of the Salish Indigenous First Nations historically heavily relied upon annual salmon runs in the Elwha basin for prosperity and spiritual wellness. They fought the dams for many decades. By joining with several environmental stakeholders they were able to finally breach the bureaucracy and power of the hydroelectric industry, and bring down the dams. The Elwha River Ecosystem and Fisheries Restoration Act of 1992 was the key legislative piece to trigger the removal of the Elwha River Dams. The Act specified that the Secretary of Interior should purchase the dams to initiate complete restoration of the Elwha River, including its anadromous fisheries.

Removal of the Elwha River dams occurred in stages to minimize the downstream disturbance of the sediment pulse, with reservoir drawdown starting in September 2011. The two dams were removed in phases, with bed material first moving past the breached 32-m Elwha Dam site in spring 2011 and across the breached 64-m Glines Canyon Dam site autumn 2011, although removal of the structures extended to 2014 (East et al., Reference East, Logan and Mastin2018). Over a five-year period (2011–2016) ~20 million tons of reservoir deposits were reworked by lateral channel migration and incision. This was about two-thirds of the total sediment volume stored in the reservoir, which resulted in downstream sediment pulses being some twenty to forty times greater than “natural” annual maximum sediment loads (East et al., Reference East, Logan and Mastin2018).

By 2015, already 90% of the reservoir deposits had been discharged to the coastal zone (Ritchie et al., Reference Ritchie, Warrick and East2018). Only minor amounts of long-term storage of reservoir sediment will occur as overbank deposits, although some sediment storage evidently occurred within side channels (East et al., Reference East, Logan and Mastin2018). The median particle size decreased from cobbles (144–174 mm) before dam removal to gravel (18 mm) by 2013. By summer 2017, particle size had become somewhat coarser to 40 mm (gravel), but was nevertheless considerably finer than before dam removal (Figure 4.41). Because the particle size did not change upstream of the dam (in the control reach), it is concluded that downstream fining of the channel bed occurred because of the reworking of reservoir deposits.

Figure 4.41.

Downstream changes to Elwha River channel before and after dam removal (photos looking upstream). Note changes to particle size of bed material and presence of wood: (A) One week prior to dam removal (September 2011). (B) Eleven months after dam removal, but prior to arrival of main sediment pulse (August 2012). (C) Shortly after sediment pulse (September 2014). (D) Almost six years after dam removal, four years after sediment pulse (July 2017).

(Source: East et al., 2018.)

Morphologically, the downstream channel reaches adjusted to the pulses by changing from a riffle morphology to a plane-bed morphology. Surveys for downstream channel cross sections showed variable geomorphic response and recovery (East et al., Reference East, Logan and Mastin2018). By spring 2013, the reach averaged bed elevation change ranged from about +0.38 m to +0.82 m of aggradation (Figure 4.42). But already within a few years most of the channel bed aggradation from the initial sediment pulse had been reworked, as channel incision ensued. By summer 2017, the average bed elevation changes at downstream research sites showed a net aggradation, but it had decreased to +0.15 to +0.35 m. These average values mask local-scale geomorphic variability, and channel bars are indeed now a more prominent component of the channel bed environment than prior to dam removal, with some bars +1.0 m. Additionally, the channel thalweg displays greater lateral mobility than previously, with one reach having laterally migrated by about 30 m.

Figure 4.42.

Reach-averaged channel-bed elevation changes between seasonal topographic surveys, referenced to earlier baseline survey data (not shown). Control reach averaged six cross sections spaced along a ~100-m channel segment located 1.5 km upstream of the upper end of Lake Mills (Glines Canyon Dam reservoir). The downstream reach was centered at 1.9 km below Elwha Dam and included six cross sections over a 172 m long channel segment. The two arrows indicate the approximate date when reservoir deposits were first transported past the two breached dam sites (Elwha, left and Glines, right).

(Source: East et al., 2018.)

Ecological changes followed geomorphic changes after removal of Elwha River dams. Tributary streams above the Elwha River Dam were being recolonized for spawning within a couple years by the river’s iconic fish, Coho salmon (Liermann et al., Reference Liermann, Pess and McHenry2017). At the coast, the copious sediment volumes resulted in about 1 m of vertical aggradation and some 400 m of lateral (deltaic) aggradation (Ritchie et al., Reference Ritchie, Warrick and East2018), which are already being colonized by pioneer vegetation (Foley et al., Reference Foley, Warrick and Ritchie2017). The prograding river mouth is associated with a salinity shift from coastal to more freshwater environments, including fish and macroinvertebrates communities. A number of activities are being conducted by government agencies and environmental organizations, including monitoring of sediment, vegetation, benthic invertebrates, including radio telemetry to map the movements of anadromous fish. The Elwha River dam removal project specifically applies to humid braided rivers (Figure 4.43), although a number of the findings are in accord with dam removal trajectories observed in other settings (Foley et al., Reference Foley, Warrick and Ritchie2017).

Figure 4.43.

Conceptual model illustrating changes in sedimentary and geomorphic regime before, during, and after dam emplacement and dam removal based on the Elwha River dam removal project in the Olympic Peninsula of Washington (United States). (A) Prior to dam emplacement. (B) Dammed river, showing deltaic reservoir sedimentation. Downstream of the dam the river is incising, narrowing, and channel bed material is becoming coarser. (C) Initial post‐dam removal phase (weeks to one to two years), showing braided channel above and below the former dam site with exposed reservoir deposits. Longitudinal profile of water surface and reservoir sediment includes migrating knickpoint. Downstream of the dam, new deposition of finer‐grained sediments buries coarse-grained armored channel bed sediment that characterized the impounded system. (D) Later in the post-dam removal phase (two to ten years), channel incision occurs through reservoir deposits and into initial post-dam removal deposits downstream of the dam.

(Figure source: East et al., 2018.)

4.4.6.1 Dam Construction and Dam Removal: Contrasting International Trends

Globally there are two contrasting international paradigms with regard to the environmental impact of dams on rivers: dam construction and dam removal. In Asia, South America, the Balkans in Europe, parts of Africa, and northern Australia, the main trend is dam construction – primarily large hydroelectric dams – but also for new irrigation projects to support agriculture and food production. In most of Europe, North America, and southern Australia, where “contemporary” dams have been installed over a longer period of time, the dominant paradigm is dam removal. While mainly limited to smaller structures, dam removal is a vital step toward improved riparian health. The large number of case studies, varying by physical geography, history, engineering, and socioeconomic context, collectively provide an invaluable database of “lessons learned” to inform removal of larger dams in the future.

The benefits of dam removal on a large river would extend far beyond the downstream limits of the river basin, and would be symbolic of a fundamental shift in human–nature relationships, and homage to the legacy of scholars that established a foundation of knowledge regarding dam impacts to riparian environments.

4.5.2.1 Huanghe River Reservoir Sedimentation: Changing Strategies

Options for managing reservoir sediment need to be adaptable to changing conditions, including climate change and land use change (Wang and Hu, Reference Wang and Hu2009; Graf et al., Reference Graf, Wohl, Sinha and Sabo2011). This can require changes to the dam structure and reservoir management strategies that differ from the original design. Such is the case with the Huanghe River basin. Reservoir sedimentation problems along the Huanghe are well known, and some 21% of the total storage capacity for its seven main-stem reservoirs (i.e., Table 4.8) was lost to reservoir sedimentation by 1989, only about two decades after the first dam was commissioned (Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017).

Sanmenxia Dam was completed in 1960 (closure) as the first main-stem dam on the Huanghe River (i.e., Table 4.8). The original design was to supply hydropower, irrigation, navigation and downstream flood control, the latter being an epic problem along the lower Huanghe associated with numerous avulsions (see Chapter 7). The initial reservoir management strategy was to store water (Table 4.19). Unfortunately, Sanmenxia was not designed to address the enormous Huanghe sediment loads, the largest sediment loads on Earth, supplied by a legacy of land mismanagement in the upstream Chinese Loess Plateau. Copious sedimentation immediately ensued after dam closure, as only a meagre ~7% of incoming sediment was transported downstream (Table 4.19). A staggering 93% of sediment was stored within the reservoir. Incredibly, by 1965 Sanmenxia reservoir had already lost 41.5% of its reservoir capacity below 335 masl and 62.9% of its storage capacity below 330 masl (Figure 4.51). The sedimentation problems resulted in having the dam redesigned to accommodate sediment sluicing beginning in the early 1970s (Table 4.19), so that the muddy water was released during high-flow events. By the mid-1970s the reservoir had already attained a more consistent storage capacity (Wang et al., Reference Wang, Wu and Wang2005).

Table 4.19Changes in reservoir management strategies of Sanmenxia Reservoir and sedimentary response

Data source: Wang et al. (Reference Wang, Wu and Wang2005).

Figure 4.51.

Longitudinal profile of reservoir deposits for Sanmenxia Dam, Huanghe River. Abrupt aggradation occurred soon after closure, and cessation of aggradation after effective reservoir management strategies implemented.

(Source: Wang et al., 2005.)

The revised sediment sluicing strategy for Sanmenxia Dam is now part of a larger sediment management approach for the Huanghe basin (Figure 4.52). Since 2002 the Yellow River Conservancy Commission regulates river flow and sediment transport to manage channel bed degradation along a cascade of dams and reservoirs (Liu et al., Reference Liu, Walling, Spreafico, Ramasmy, Thulstrop and Mishra2017; Ma et al., Reference Ma, Nittrouer and Naito2017; Kong et al., Reference Kong, Latrubesse, Miaoa and Zhou2020). The importance of developing an appropriate reservoir sediment management strategy for the Huanghe is as important as any river in the world. Flood disasters along the lower Huanghe River are linked to channel bed aggradation attributed to landscape and sediment mismanagement. The substantial population and economic activities associated with the lower Huanghe riparian environments means that the issue is not only a vital societal concern, but it also influences fluvial dynamics in the lower basin related to channel avulsion, and has implications to the global economy (Section 7.5.2).

Figure 4.52.

Sediment flushing event of Sanmenxia Dam, Huanghe River. Sanmenxia Dam closed in 1960 and is the first main-stem dam on the Huanghe River (contributing drainage area 688,400 km²).

(Photo date: July 26, 2013. Source: R. Mueller, licensed by CC.)

4.5.2.2 Closure: Reservoir Management and Sustaining Rivers

The appropriate sediment management approach for a particular dam and river basin is dependent upon a variety of factors including sediment volume and sediment type (sand/gravel or silt/clay), reservoir geometry (size and shape), streamflow variability, downstream geomorphic and environmental conditions, and expenditures. Adapting the appropriate sediment management strategy can extend the life span of reservoirs beyond the conventional design-life paradigm (Figure 4.53). And, as noted along the Huanghe River, sediment management approaches should be adaptive to accommodate changing boundary conditions and societal priorities because of its impact on the supply of upstream sediment loads (Graf et al., Reference Graf, Wohl, Sinha and Sabo2011; Kondolf et al., 2014a). Identification of an appropriate reservoir sediment management strategy should be seen as a fundamental component of integrated river basin management to support navigation and associated economic activities, mitigate against climate and land use change, and to sustain downstream riparian and linked coastal environments.

Figure 4.53.

Models of dam and reservoir design related to sediment management. Conventional “design-life” paradigm for dams and reservoirs can be modified toward a sustainable use paradigm. New dam construction should include features adaptable to changing sediment management strategies.

(Source: Annandale et al., 2016.)

It is interesting to ponder a sediment management strategy to Gavins Point Dam (Figure 4.21) on the Missouri River in view of the need to provide sediment to support physical riparian habitat and related biodiversity (NRC, 2011), and the urgent need to supply sediment to construct wetlands in the Mississippi delta (Ahn et al., Reference Ahn, Yang, Boyd, Pridal and Remus2013; Kemp et al., Reference Kemp, Willson, Rogers and Binselam2014). Proposed drawdown flushing strategies designed to scour reservoir deposits and mobilize stored sediment for downstream transport, however, are complicated by the need to consider both the sediment budget of the lower Mississippi as well as specific habitat needs of aquatic species along the lower Missouri, including the endangered Pallid Sturgeon (Scaphirhynchus albus) (Coker et al., Reference Coker, Hotchkiss and Johnson2009; Elliot et al., Reference Elliott, DeLonay, Chojnacki and Jacobson2020; Jacobson et al., Reference Jacobson, Blevins, Bitner, James, Rathburn and Whittecar2009). Nevertheless, identifying a “sustainable” approach to mobilize reservoir sedimentary deposits stored behind Gavins Point Dam, and other Missouri basin dams (Table 4.11), should be of highest priority to responsible government organizations.