Introduction

Eutrophication is defined as the nutrient over-enrichment of waterbodies that results in high rates of primary production (Paerl et al. 2001). Eutrophic conditions can occur naturally, depending on such factors as watershed–lake ratio, waterbody volume, water residence time, and the nutrient status of local bedrock and soils. Cultural eutrophication, on the other hand, is caused by human activities in the watershed and has become widespread, threatening water quality globally (Njagi et al. 2022). For example, economically important waterbodies around the world, including Lake Victoria in Africa, Lake Kasumigaura in Japan, Lake Taihu in China, and Lake Erie, on the USA-Canada border, are affected by cultural eutrophication (Nwankwegu et al. 2019). The magnitude of ecological and economic negative impacts of eutrophication demands global prioritization of its prevention, management, and control (Qin et al. 2013).

The primary strategy used to prevent eutrophication is to restrict external nutrient inputs to waterbodies (Fig. 1) (Harding and Paxton 2001; Paerl et al. 2016). Whereas the approach appears straightforward, it is in fact tedious, challenging, time-consuming and expensive. This is, in part, because non-point nutrient sources, for example runoff from agriculture, are difficult to control. Additionally, nutrients such as nitrogen may be deposited from the atmosphere and originate from remote sources, making their control almost impossible (Mayanga et al. 2014). This was reportedly the case at Lake Victoria where atmospheric deposition contributed 90% of phosphorus and 94% of nitrogen inputs into the lake (Nyenje et al. 2010). Furthermore, human activities generate large amounts of nutrients that eventually get discharged into waterbodies. As the global human population has grown, anthropogenic nutrient sources have become more diverse, intensified, and more challenging to control (Dube et al. 2017). In addition, nutrients accumulated in the waterbody sediments can be released back into the water column under some physico-chemical conditions, where they support greater primary productivity and counteract or delay the effects of external nutrient-load reductions (Hye et al. 2019). For example, in Lake Erie, phosphorus loading from the sediments is equivalent to 8–20% of the external nutrient load (Paytan et al. 2017). Similarly, in Lake Pontchartrain (Louisiana, USA), the internal phosphorus load accounts for approximately 30–44% of the total annual phosphorus load to the waterbody (Roy et al. 2012). Consequently, tangible benefits of external nutrient load reduction to a waterbody may take too long to manifest, whilst incurring huge costs (Jeppesen et al. 2011). For instance, at Lake Taihu, US $16.25 million was spent on nutrient control in 2007, but by 2013 nutrient load reductions had not yielded positive results and a massive cyanobacteria bloom occurred in the summer of that year (Merel et al. 2013). Therefore, costs may limit the application of external nutrient restriction programs in some areas, especially in low-income countries. Furthermore, the benefits of external nutrient load reductions can also be compromised by the effects of climate change on the watershed hydrology and nutrient loading dynamics, water temperature, lake mixing regime, and internal nutrient dynamics (Jeppesen et al. 2011).

Fig. 1
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Strategies for the prevention, management, and control of eutrophication in surface water bodies

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Programs to restrict external nutrient loading into water bodies have often produced limited success, which therefore, necessitates that eutrophication control and management include within-waterbody remediation techniques, to manage internal nutrients already accumulated in the water body and suppress the symptoms of eutrophication (Fig. 1) (Lürling and Mucci 2020). Within-waterbody remediation strategies include chemical, physical, or biological interventions designed to reduce water-column nutrient concentrations and ameliorate the symptoms of eutrophication (van Ginkel 2011; Harding 2015; Harding et al. 2009). These techniques have different modes of action and present both advantages and disadvantages with respect to capital and operation costs, as well as ecosystem impacts. Their effectiveness depends on many factors, such as external nutrient loading rate, water body morphometry, internal nutrient inventory, water-column stratification/circulation patterns, sediment quality, flushing rate, and phytoplankton diversity (Lürling and Mucci 2020). The most effective remediation strategies are designed on a case-by-case basis in accordance with the waterbody’s morphometry, limnological variables, extent of nutrient enrichment, and water uses. The initial planning steps are key to establishing cause-and-effect linkages between water quality concerns and proposed remediation techniques (Kibuye et al. 2020). Additionally, because each remediation technique affects multiple components of aquatic ecosystems, including water quality, appropriate diagnostics are necessary to identify the mechanisms and trajectory of restoration, as well as possible inadvertent effects (Hickey and Gibbs 2009). Consultations with relevant stakeholders, including affected communities, about the anticipated consequences of different proposed treatment techniques are integral to successful within-waterbody remediation (Dennis and Dennis 2019).

This work reviews the eutrophication status of inland water bodies in South Africa (SA). The major drivers of eutrophication, relevant within the SA context, are identified and nutrient loading patterns through time are presented for some water impoundments. Furthermore, the socioeconomic impacts of eutrophication are detailed to highlight their effects on the broader SA economy and the everyday lives of South Africans. In the last segment of the review, the different measures undertaken to prevent, manage and control eutrophication are reviewed, with a discussion of their success rates, limitations, potential opportunities, and lessons learned.

Eutrophication and Cyanobacterial Blooms in South African Inland Waters

South Africa has few natural freshwater lakes, most of which are distributed in remote areas and not easily accessible or useful for urban water supply. Therefore, exploitable water in SA is limited to rivers and groundwater. To meet the high demand for potable, industrial and irrigation water, and to ensure a constant supply from the irregularly flowing rivers, dams have been constructed on all the major rivers, creating artificial freshwater impoundments (Oberholster and Ashton 2008). Currently, there are approximately 4755 large reservoirs each with a capacity of more than 50 000 m3, and a wall height of > 5 m (Stevens and Van Koppen 2015).

Eutrophication of freshwater impoundments is identified as the major form of surface-water pollution in SA, first reported in the early 1960s (Rasifudi et al. 2023). Several surveys have been conducted over the years to assess its intensity and geographic distribution. For example, a survey was conducted between 1985 and 1999 in 48 impoundments from 12 river catchments, covering all nine provinces. Eutrophication indicators, including total phosphorus (TP), chlorophyll-a (chl-a), total nitrogen to total phosphorus ratio (TN:TP) and cyanobacterial dominance, revealed that 23% of the impoundments were hypereutrophic, 23% eutrophic, 33% mesotrophic and 17% oligotrophic (van Ginkel 2002). The eutrophic and hypereutrophic water impoundments were spatially distributed across the country, with Gauteng Province leading in the number of such nutrient-rich impoundments. A more comprehensive study was conducted between 2002 and 2012 using remote satellite imaging of 50 impoundments with large surface areas. The data confirmed that nutrient enrichment in major impoundments was widespread and severe, with 72% of the dammed waterbodies being hypereutrophic, 6% eutrophic, 8% mesotrophic and 14% oligotrophic (Matthews 2014). In another survey, in data were collected by the National Eutrophication Management Program (NEMP) for 393 dammed impoundments. In the period between 2016 and 2020, there was a 4%, 2% and 9% increase in the number of hypereutrophic, eutrophic, and mesotrophic waterbodies, respectively. At the same time, the number of oligotrophic waterbodies declined by 14% (Rasifudi et al. 2023).

In earlier eutrophication surveys, rivers were generally not included, in part because some SA river systems contain high concentrations of suspended solids and are therefore turbid, making it difficult to establish a chlorophyll-nutrient relationship (Oberholster and Ashton 2008). Only after reports of the symptoms of eutrophication became more pronounced river systems, were they included in surveys. According to the NEMP, 18 of 25 major river systems are eutrophic. These include the Olifants, Orange (Mpumalanga), Crocodile, Pienaars, Vaal, Jukskei (Gauteng), Buffalo, Berg, Umgeni, Umsunduzi (KZN Province), and Modder (Northern Cape Province) Rivers (Griffin et al. 2014).

All the studies indicate that eutrophication in SA is widespread in the 9 provinces and shows an increasing trend in its intensity. This trend is further corroborated by the first appearance of eutrophication symptoms in dammed waterbodies that had not previously shown them. One example is the Vaal Dam impoundment which has been in the spotlight lately because of concerns that its water quality is progressively deteriorating, raising fears that the potable water supply may be affected negatively (Ndlela et al. 2016; Ntshalintshali 2019). In the next section, some of the key drivers of eutrophication in SA are identified and discussed.

Key Drivers of Eutrophication

Poorly treated effluents discharged from wastewater treatment plants (WWTPs) and industry, runoff from agricultural activities, untreated sewage from leaking and overflowing sewer systems, as well as runoff from informal settlements are some of the primary drivers of eutrophication in SA (Fig. 2) (Department of Water and Sanitation 2022b). The climate conditions in SA are also favorable for eutrophication as most parts of SA are arid to semi-arid and characterized by high temperatures and erratic, unpredictable weather extremes during the wet and dry seasons (Molobela 2011; Petterson 2019a). During episodes of high rainfall, nutrients are washed into water impoundments, whereas during the dry season, rates of inflow and outflow decline, evaporation rates increase, and nutrients become concentrated in the water column (Oberholster and Ashton 2008). The Department of Water and Sanitation (DWS) notes that SA has several well-crafted national laws and regulations designed to protect water resources and control effluent discharge. These include the National Water Act of 1998 (NWA) and the Water services Act of 1997 (WSA) which are both legal frameworks for water resources governance. Both legislations place emphasis on the management of the entire catchment and use of freshwater without negatively affecting the aquatic ecosystem (Department of Water and Sanitation 2022a). However, there is a lack of, or poor enforcement of these laws which is recognized as another factor that contributes to intensified eutrophication of waterbodies.

Fig. 2
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A schematic presentation showing the primary drivers of, and symptoms of eutrophication in SA inland waters

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Wastewater Treatment Plants

In some SA water impoundments, discharges from WWTPs account for > 50% of the total nutrient load (Germishuys and Diamond 2022; Zulu 2020) making them the major drivers of nutrient enrichment (Griffin et al. 2014; Joshua Rasifudi et al. 2023). In addition to poorly treated effluent discharge from WWTPs, raw sewage spills from leaking and overflowing sewer systems and untreated discharge from informal settlements are sources of nutrients in waterbodies. The Green Drop (GD) reports, in which WWTP’s are evaluated in terms of operating capacity, skills availability, delivery, and functionality depicts the poor state of wastewater treatment services in SA (Table 1) (Department of Water and Sanitation 2022c). The latest report released in 2022 indicates that out of 995 plants evaluated only 2% obtained GD certificates (functioned efficiently) and 34% were found to be in a critical state (Department of Water and Sanitation 2022c). The poor state of wastewater treatment services in SA is further evident from the frequency of raw sewage spills in municipalities across the country, which prompts communities to report such events to the South African Human Rights Commission (SAHRC). As a result, the SAHRC confirmed that many wastewater treatment plants were dilapidated and recommended that the poor state of wastewater treatment in SA be declared a national state of disaster (South African Human Rights Commission 2021). The main challenges facing wastewater treatment services are system overload, deteriorated sewage collection and treatment infrastructure, shortage of skilled personnel, poor management, general unsatisfactory performance of municipalities, and vandalism (Ntombela et al. 2016; Oberholster and Ashton 2008). Currently the country is facing an energy crisis which negatively impacts wastewater treatment services as well. Therefore, it is anticipated that the amount of raw and poorly treated wastewater being discharged into surface waterbodies will increase dramatically. Nonetheless, the wastewater treatment challenges are not unique to SA as the United Nations (UN) laments that inadequate wastewater treatment is a challenge facing both low-income and high-income countries. In low-income countries, approximately 10% of wastewater undergoes any form of treatment compared to 70% in high-income countries (United Nations World Water Assessment Programme 2017).

Table 1 Green drop performance trends of WWTPs in SA between 2009 and 2021 (Department of Water and Sanitation 2022a)
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Agricultural Runoff

Historically, the contribution of agricultural runoff (crop production and livestock farming) to nutrient loading in waterbodies has been masked by that of WWTPs and a lack of methods for proper quantification (Xia et al. 2020). However, reports from countries such as the United States of America, Europe and Australia point to the substantial contribution of agriculture towards nitrogen and phosphorus loading in water impoundments (Harrison et al. 2019). For instance, in Lake Taihu, China, agricultural runoff contributes 52% of the nitrogen load and 54% of the phosphorus load. Similarly, in Italy the estimated proportions of nutrient loads to different waterbodies from agriculture are 24% for TN and 71% for TP (Sun et al. 2012). In SA, valuable arable land is mostly found along riverbanks (Van Der Laan and Franke 2019). The DWS has identified several poor agricultural practices that are prevalent in SA, including excessive fertilizer and pesticide application, improper tillage, exorbitant irrigation, and encroachment upon and destruction of riparian zones and wetlands (Van Der Laan and Franke 2019). All these practices are known to increase the amount of agricultural runoff that reaches fresh waterbodies. Furthermore, a survey conducted in the Middle Olifants catchment revealed that agricultural activities contribute substantially to phosphorus and nitrogen loads into nearby waterbodies, including the Loskop and Flag Boshielo impoundments and the Olifants River. In the Vaal River catchment as well, intensive fertilizer use for potato production was shown to contribute to nitrogen loading into subsurface and surface waters (Graham et al. 2012). These studies point towards a significant contribution of run off from agricultural activities to nutrient loading in inland water bodies. However, there is a need for further studies, to better quantify nutrient pollution from agricultural activities. This is necessary to inform best management practices for minimizing agricultural runoff. This is crucial because nutrient loading from agricultural runoff is on the rise as the human population continues to grow, increasing the demand for food production (Timsina 2018).

Urbanization

In the process of urbanization, the natural landscape of an area is tampered with, ultimately disturbing/changing the natural drainage systems. Urbanised areas are characterised by impervious surfaces and asphalt which prevents soil recharge by the natural cycle. Man-made structures such as roads, storm drains, all channel nutrient-rich water to nearby waterbodies and contribute to water quality deterioration (Deng et al. 2015). Urban areas are also characterised by uncontrolled population growth which increases the amount of nutrient-rich discharge reaching urban waterbodies (Oliver et al. 2019). Population increase also puts pressure on WWTP’s which then fail to efficiently treat wastewater. There is also a high demand for land to be used for agriculture and human settlement, often causing destruction and encroachment of nutrient-filtering sites such as riparian zones, wetlands, and riverbanks. These areas, therefore, lose their nutrient-filtering capacity (Dube et al. 2017). Furthermore, the pressure from uncontrolled urban population growth results in the mushrooming of informal settlements (slums) which often have high population densities and inadequate sanitation services. Therefore, the nutrient loads produced in these settlements are often very high and result in increased nutrient fluxes in urban water impoundments (Nyenje et al. 2010). In SA, the contribution of urbanization towards the eutrophication of water bodies may be severe because metropolitan areas are in river catchments and the rivers serve a dual purpose for these often densely populated areas, as sites of water supply and waste disposal. Therefore, the waste collected in river systems is transported downstream into water storage impoundments where it is sequestered (Maavara et al. 2015). As a result, urban water bodies in SA are often the most nutrient-enriched. This is the case in the highly urbanised and densely populated Gauteng Province, where all the waterbodies are eutrophic to hypereutrophic (CSIR 2010; Matthews 2014).

Symptoms of Eutrophication

In fresh waterbodies, the most immediate and detrimental symptom of eutrophication is proliferation of cyanobacterial blooms (Paerl and Otten 2013). Under conditions of excessive nutrient enrichment, cyanobacteria grow rapidly and form blooms which are dense cell aggregations (≥ 10,000 cells mL−1) that can appear as thick paint-like scums on the water surface (Harding and Paxton 2001; Sciuto and Moro 2015). In SA water impoundments and river systems, cyanobacterial blooms are the primary symptom/sign of eutrophication. In the surveillance study undertaken in 2014, cyanobacterial blooms were reported in all 50 studied impoundments. In five of the waterbodies the blooms were extensive and covered 30% of the surface area, whereas in 18, coverage was intermediate (10–30% of the surface) (Matthews 2014). In most water impoundments, cyanobacterial blooms proliferate periodically during the warm spring and summer months, whereas in a few water bodies these occur continuously, year-round. The most common genera of cyanobacteria in SA surface waters are Microcystis, Anabaena, Cylindrospermopsis, Oscillatoria, Lyngbya, Merismopedia, Pseudoanabaena, Chroococcus, Nostoc, Nodularia, and Spirulina (van Ginkel 2004). Whereas cyanobacterial blooms are the most common, there have been cases of blooms of other phytoplankton, including the dinoflagellate Ceratium hirundinella and the green alga Scenedesmus sp. (van Ginkel, 2011a). Oberholster et al. (2009) observed that the spatial distribution of cyanobacterial blooms has expanded, and the frequency and magnitude has intensified (Oberholster et al. 2009). According to the authors, climate change is the most plausible cause for the observed trend. A typical example is the Hartbeespoort Dam impoundment, in which new phytoplankton taxa were identified, including Cylindrospermopsis curvispora, Sphaerospermopsis aphanizomenoides, Sphaerospermopsis reniformis, Raphidiopsis mediterranea and Raphidiopsis curvata (Ballot et al. 2014; O’Neil et al. 2012). In the Rietvlei Dam impoundment as well, there has been a shift in the phytoplankton diversity where blooms previously dominated by Ceratium hirundinella (van Ginkel 2004) are now dominated by Microcystis sp. (Mbiza 2014; van Ginkel 2004).

The genera of cyanobacteria prevalent in 99% of SA water impoundments are mostly toxigenic and produce a range of cyanotoxins (Oberholster & Botha 2007). Microcystins, produced by Microcystis, Anabaena, Nostoc, Planktothrix, and Phormidium spp., are the most frequently quantified toxins in SA fresh waterbodies. Various microcystin congeners have been identified in different water bodies such as Roodeplaat, Rietvlei, Hartbeespoort, Vaal, Loskop, Midmar and the Limpopo Basin (Joshua Rasifudi et al. 2023). At Hartbeespoort alone, more than 41 microcystin congeners, including MC-RR, MC-LR, MC-YR, MC-WR, MC-FR, MC-LA and MC-YA were identified (Ballot et al. 2014; Mbukwa et al. 2012; Eguzozie et al. 2016; Conradie and Barnard 2012). The total microcystin concentrations quantified in SA water impoundments are often extremely high. For example, at some point the average microcystin concentrations quantified in five impoundments in the Gauteng Province (Hartbeespoort, Roodeplaat, Rietvlei, Bon Accord, and Bloemhof) were > 10,000 µg L−1. At Hartbeespoort as well, concentrations of microcystins as high as 14,400 μg L−1 have been reported (Ballot et al. 2014). These concentrations are relatively higher than those recently reported in some African countries like Tanzania (0.4–13 µg L−1), Zimbabwe (1.6–22 µg L−1), Mozambique (0.9–7.8 µg L−1), Egypt (300–877 µg L−1), Tunisia (0.01–5.37), Ethiopia (0.6–1547 µg L−1), and Ghana (0.03–3.2 µg L−1) (Chia et al. 2022).

In addition to cyanobacteria and their toxins, invasive aquatic weeds are also a common symptom of eutrophication in SA impoundments. The five most common aquatic weed species in eutrophic fresh waterbodies in SA (referred to as “the big bad five”) are: Eichhornia crassipes (water hyacinth), Azolla spp. (red water fern), Pistia stratiotes (water lettuce), Salvinia molesta (Kariba weed), and Myriophyllum aquaticum (parrot’s feather) (Hill and Coetzee 2017). Of these, hyacinth is the most problematic and widespread aquatic weed associated with eutrophication acknowledged as the second most common symptom of eutrophication (Auchterlonie et al. 2021; Germishuys and Diamond 2022; Petterson 2019b; Thamaga and Dube 2018). Their occurrence has been reported in several water impoundments including Roodeplaat and Hartbeespoort, and in the upper Vaal River, where they occur together with cyanobacteria, or alone (Coetzee and Hill 2012). With their rapid vegetative growth and high reproduction rates, hyacinths can cover a large proportion of a waterbody surface in a short span of time. The presence of abundant hyacinths in a waterbody can magnify the symptoms of eutrophication by leading to oxygen depletion, increased pH, and turbidity in the water column.

Impacts of Eutrophication on Different Sectors

In addition to disrupting the ecology of aquatic ecosystems, eutrophication has detrimental economic and social impacts that affect human lives directly or indirectly (Fig. 3). In SA, tourism, agriculture, and potable water supply are some of the sectors negatively impacted by eutrophication.

Fig. 3
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Schematic drawing showing the symptoms of eutrophication in freshwater bodies as well as their impacts on different sectors of the economy

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Recreation and Sports

The appearance of cyanobacteria blooms, invasive plants, cyanotoxins, and odorous compounds in eutrophic waters lowers their recreational and amenity values. Invasive plants and cyanobacterial blooms often block waterways and limit the extent to which waterbodies can be used for recreational activities such as swimming, fishing, and boating (Havens 2008). Cyanotoxins and malodorous compounds in water are deterrents to potential users of waterbodies and can present health risks. In a survey conducted at the Vaal and the Bloemhof impoundments, 68% and 45% of visitors, respectively, raised concerns about inferior water quality and expressed a lack of interest in visiting the waterbodies (Graham et al. 2012). Eutrophication impacts are more apparent in waterbodies that host popular sporting events, as these are sometimes cancelled because of cyanobacterial blooms or invasive plants. For instance, at Roodeplaat, events hosted by the Roodeplaat Rowing Club and Rowing SA have been negatively impacted, as the waterbody occasionally experiences blooms of cyanobacteria and hyacinth coverage. In 2019, the national Olympic squad was forced to abandon Roodeplaat and travel to Katswe, in Lesotho, to train. Additionally, the Roodeplaat Rowing Club has cancelled several national competitions because of hyacinth and cyanobacterial blooms. The club further spent close to R250,000 (US $13,000) replacing the rowing course, which was damaged (Bega 2021). Rowing SA is scheduled to host the 2023 Olympic rowing competition. However, the competition will be cancelled if the organization cannot remove hyacinth plants covering 65% of the impoundment surface, resulting in the loss of money spent during the bidding process. In another case, the annual Dusi Canoe Marathon, which attracts > 2000 paddlers and is held at the Msundusi River in the KwaZulu Natal Province, is under threat because of the eutrophic state of the river. On several occasions, cases of ill health were reported by paddlers, and these were attributed to the poor quality of the eutrophic waters. If cancelled, individuals and companies who provide services such as food and accommodation during these events will be economically affected by loss of income (Frost and Sullivan 2010).

Agriculture

In the agriculture sector, the negative impacts of eutrophication are felt mostly in aquaculture, crop irrigation, and livestock farming. If eutrophic water with high concentrations of cyanotoxins and T&O compounds is used in aquaculture farms, fish become tainted with these cyanobacteria secondary metabolites and lose their commercial value (Paerl and Otten 2013). For example, annual monetary losses incurred because of eutrophication were estimated at R12,000,000 (US $624,000) in trout production farms in the Mpumalanga and Western Cape Provinces (Frost and Sullivan 2010). Also, the livelihoods of communities that rely on fish caught from local waterbodies as a cheap source of protein are compromised by eutrophication. In eutrophic waterbodies, desirable fish species with high nutrient value are often replaced by non-desirable species that are less palatable. To aggravate the situation, eutrophic waterbodies experience frequent massive fish kills because of anoxia or intoxication, which reduces harvestable fish stocks. This has been the case at Juskei River, Vaal, Roodeplaat and Hartbeespoort impoundments (Dube et al. 2017).

With respect to irrigation, which uses more than 62% of the country’s water resources, the impacts of eutrophication manifest through crop plants absorbing cyanotoxins, algae cells and weeds blocking irrigation channels and fouling equipment, as well as livestock deaths caused by cyanotoxin poisoning (Pindihama and Gitari 2019). When plants absorb cyanotoxins from irrigation water, their growth rate and quality are compromised. Toxins absorbed by crop plants present a possible health risk to consumers. The quality of produce may be reduced to levels below export standards, resulting in the loss of international export markets (Melaram et al. 2022). For example, tobacco plants irrigated with eutrophic water from the Hartbeespoort impoundment absorbed T&O compounds which affected the taste and commercial value of the tobacco. Farmers experienced economic losses, as the tobacco could not be exported and the farming project was eventually abandoned (De Lange et al. 2016; Pindihama and Gitari 2019). Similarly, local suppliers of horticultural products have stopped sourcing products from farmers around the Hartbeespoort impoundment, fearing product contamination. This has resulted in loss of income (du Preez et al. 2018; Pindihama and Gitari 2019). The irrigation systems are also affected by filamentous algae cells and aquatic weeds in irrigation water, which block irrigation canals and clog components of irrigation systems such as sprinklers and drip irrigation devices. As a result, farmers bear inflated operation costs from having to clear blocked canals, replace clogged parts and remove algal cells and cyanotoxins from irrigation water (Pindihama and Gitari 2019). For example, wheat and citrus farming in the Middle Olifants Irrigation Scheme incurs costs of R2890 (US $150) per hectare, amounting to a total of R48.6 million (US $2.53 million) per year, associated with the treatment of eutrophic irrigation water (Mudaly and van der Laan 2020). Another impact of eutrophication in the agriculture sector is the death of livestock from exposure to cyanotoxins. An estimated 60% of farm ponds in SA are affected by cyanobacterial blooms at some point during the year and livestock deaths are prevalent in those areas. In those waters, cyanobacterial cells are often concentrated near the shores where animals drink. When cells die, cyanotoxins are released into the water, where they remain dissolved for extended periods. When animals ingest cyanobacterial cells or dissolved cyanotoxins, they die from acute intoxication, often in large numbers. The most notable case of livestock cyanobacteria poisoning was the death of 400 dairy cows in the Kareedouw District of Tsitsikammma (Frost and Sullivan 2010). At Kruger National Park, large numbers of rhinoceroses, zebras, elephants, and wildebeest have died from microcystin intoxication (Buss and Bengis 2012). At times, intoxication cases associated with microcystins have ranked as the fifth and tenth most important in Gauteng and Mpumalanga, respectively (Frost and Sullivan 2010).

In the year 2022 agriculture contributed 2.6% to the gross domestic product (GDP) and an estimated 5% to formal employment. So, the negative impacts from eutrophication will affect the country’s GDP and contribute to job losses. Most importantly negative impacts in the agricultural sector directly pose a threat to food security (Stevens and Van Koppen 2015).

Potable Water Supply

Cyanobacterial cells, cyanotoxins, and T&O compounds, as well as high nutrient concentrations and pH levels in eutrophic waters reduce the production capacity of water treatment plants (WTPs), whilst increasing production costs. Cyanobacterial cells present in raw water clog filter units, necessitating frequent filter replacement and backwash, which reduces production and increases operation costs (Barnard et al. 2014). Some species of cyanobacteria such as Oscillatoria, Planktothrix and Anabaena can pass through the filtration apparatus and release more toxins into the water further down the treatment train if the cells rupture. Other cyanobacteria metabolites (mucopolysaccharides) can chelate metal ions added as coagulants in water treatment, which limits the flocculation process and increases concentrations of metal ions in the water, creating a health hazard for consumers (Pivokonsky et al. 2016). In many instances when WTP’s are affected by eutrophication potable water supply is disrupted, which contributes to water scarcity and negatively impacts on human health as it increases the risks of water-borne disease outbreaks.

To minimize the negative impacts of eutrophication in the potable water supply, several WTP have been upgraded to include advanced treatment technologies with improved efficiencies in removing cyanobacteria secondary metabolites. That has been the case in several water treatment plants, including the Rietvlei, Roodeplaat, Brits, and Schoemansville WTPs (Roux et al. 2010). At times when there were cyanobacterial blooms at the Rietvlei impoundment, the Rietvlei WTP, which gets its raw water from Rietvlei, experienced fluctuations in the raw water quality in terms of pH, turbidity, and T&O compounds. These necessitated alterations to the water treatment chemicals used as well as their dosage (van Vuuren 2012). The plant output was reduced from 40 to 20 megaliters per day, as filters clogged with cyanobacterial cells. Therefore, the original treatment stream of flocculation, settling, filtration, and chlorination was upgraded to include air flotation, incorporated in the 1980s (Haarhoff and van Beek 1996), followed by addition of granular activated carbon in the 2000s and an ozonation unit (Barnard et al. 2014). Similarly, at the Roodeplaat WTP, production was impaired by the highly eutrophic raw water sourced from Roodeplaat impoundment, prompting an upgrade of the ozonation and granular activated carbon units to cope with the high algal toxins and T&O compounds (van Schalkwyk 2013). At the Schoemansville WTP, production capacity was reduced from 15 to 10 megaliters per day during blooms of cyanobacteria. As eutrophication at Hartbeespoort reached alarming levels, the plant struggled to reduce microcystins to acceptable levels. At some points, microcystins more than the recommended guideline of 1 µg L−1 were detected in treated potable water supplies. Operations in the WTP were suspended and affected residents were supplied with water sourced from alternative safe supplies. It was then that the plant was upgraded to include activated carbon and ozone as a polishing step to remove soluble microcystins (Mbiza 2014). The Brits WTP also struggled to reduce T&O compounds in its water to acceptable levels and consumers resorted to drinking bottled water. Subsequently, the plant was upgraded to increase production capacity by 20 megaliters per day, and dissolved air flotation, sand filtration, ozonation, lime addition as well as chlorination units were installed (Roux and Oelofse 2010). Despite the upgrade, seven of the 14 sand filter units were recently blocked by high algal cell densities in raw water, reducing the plant output dramatically. Consequently, water supplies were disrupted for several days and residents were supplied with water from alternative sources (Britspos 2023).

It is worth noting that the upgrades in water treatment plants demand huge capital costs and subsequent increases in operational costs. For instance, upgrades at the Roodeplaat and Schoemansville WTPs amounted to R7,695,697 (US $400,000) and R250,000,000 (US $13,000,000), respectively. Often these technological upgrades are not aimed at increasing the production capacity, but rather at maintaining existing production levels and water quality. However, if the challenges of water quality persist, there is no guarantee that there will be no need for further upgrades in the future. Therefore, there is a need for to put more efforts towards the prevention of eutrophication as opposed to the treatment of its symptoms. In their study, Roux and Oelofse (2010) alluded to the fact that prevention of eutrophication is economically viable when compared to the costs of treating eutrophication symptoms.

Economy

Economic impacts of eutrophication are associated with increased costs of water treatment and losses in the tourism and agriculture sectors. In the United Kingdom, annual economic impacts of eutrophication were estimated at US $105–160 million in 2003, and US $2.2 billion in the USA in 2008 (Dodds et al. 2009; Pretty et al. 2003). In SA, there is a shortage of studies that have quantified the economic impacts of eutrophication. Nonetheless, a survey was undertaken to develop a generic model for use in quantifying the cost of eutrophication in agriculture, potable water supply and property development in SA. Using the developed model, significant and meaningful relationships between nutrient or chlorophyll concentrations and water treatment costs were established at the Zuikerbosch and Balkfontein WTP’s. It was shown that a 1% increase in N concentrations results in a 0.2% and 0.3% increase in water treatment cost at the Zuikerbosch and Balkfontein WTPs, respectively. In agriculture, increased nitrogen concentrations in irrigation water were also statistically correlated with the costs of production (Graham et al. 2012). These preliminary results prove that eutrophication is an economic burden threatening economic sustainability and development in SA. Based on the costs reported in other countries, it can be predicted that economic losses caused by eutrophication in SA could amount to hundreds of millions of Rands per year. There is therefore an urgent need to quantify the economic impacts of eutrophication in SA to fully appreciate the need for eutrophication control (De Lange et al. 2016).

Prevention, Management and Control of Eutrophication in South Africa

Surveillance, Monitoring and Reporting

Proper surveillance, monitoring and reporting are critical components for prevention and management of eutrophication. The NEMP was the first and major surveillance program initiated by DWS in 2002 to monitor and report on eutrophication and its associated problems in major water impoundments across the country (Chudleigh et al. 2000). In this program, water quality monitoring variables including nutrient (N & P) and chlorophyll-a concentrations, phytoplankton diversity, and Secchi disk depth are measured from grab samples collected bi-weekly from major impoundments and rivers. The data are then made publicly available. However, the sampling methods used are labor-intensive and hence the program does not cover all water impoundments in SA. Furthermore, the available data have a lot of gaps (Joshua Rasifudi et al. 2023). Recent advances in remote satellite imaging have made it possible to incorporate this technology into the NEMP as a surveillance tool, to provide regular monitoring and reliable data (Department of Water and Sanitation 2022b). Efforts at developing and optimizing satellite imaging algorithms have culminated in the development of a website where data on major impoundments are available in real time (Matthews 2022).

Prevention

The 1.0 mg L−1 Orthophosphate Standard

The special ortho-phosphate standard of 1.0 mg L−1 discharged wastewater effluent was a eutrophication preventive measure piloted in WWTPs that feed into eight major water impoundments (Chutter 1989). These were the Rietvlei, Hartbeespoort, Roodeplaat, Klipvoor, Bon Accord (in the Crocodile River catchment in Gauteng and Northwest Province), Inanda (in uMgeni catchment in Kwazulu Natal Province), Shongweni (in Mlaas catchment in Kwazulu Natal Province) and the Laing (in the Buffalo catchment in Eastern Cape Province). The standard was specifically implemented in WWTPs because they were identified as the primary contributors to nutrient loads entering water impoundments. The focus was on phosphates because it is accepted that P is generally the limiting nutrient that controls cyanobacteria productivity in fresh waters (Paerl et al. 2011; Schindler et al. 2016).

After implementation of the standard the water impoundments responded differently, with significant reductions in phosphate concentrations observed at the Bon Accord, Hartbeespoort and Rietvlei impoundments. At Hartbeespoort, this was accompanied by a reduction in the formation of “hyperscums” (Chutter 1989). In the Laing, Klipvoor, Inanda and Shongweni impoundments, the implementation of the phosphate standard had no effect, i.e., there were no significant reductions in phosphate concentrations. Roodeplaat was an exception, in that there was a significant increase in nutrient concentrations after the implementation of the standard. A simple explanation to this could be that the implementation of the standard coincided with an increase in the rate of nutrient loading into the impoundment. Even in the water impoundments where there were reductions in phosphate concentrations, these changes were not sufficient to change their eutrophic status, implying that the introduced P restrictions were ineffective (Griffin et al. 2014). This stands in stark contrast to reports of significant reductions in bloom occurrence in waterbodies such as Lake Erie, Lake Washington, and the Potomac River in the USA, where similar P restrictions were introduced (Merel et al. 2013). Nonetheless, there are other reports on the failure of P restrictions to prevent eutrophication, which suggest that P restrictions in wastewater may be hard to achieve and may not provide a practical solution for eutrophication control (Chorus et al. 2020).

In SA, the limited success of external P restrictions was attributed to several inadequacies in how the standard was implemented. First, a ‘blanket’ approach was adopted in which the eutrophication threshold for each waterbody was not taken into consideration. The eutrophication threshold is crucial as it considers the trophic status of the waterbody and measures the practical levels of nutrient reductions that are necessary to achieve an effective shift in the trophic status (Harding 2008). The standard also targeted only WWTP effluents and ignored other external sources of nutrients that contribute to eutrophication (Chen et al. 2017). Nonetheless, the failure of WWTPs to comply with the restrictions was the major factor contributing to the failure of this special standard. For example, in a survey conducted at the Baviaanspoort WWTP early in the implementation of the standard, 95% of collected samples exceeded the 1 mg L−1 phosphate concentration, with 52% recording > 5 mg L−1 and 19% > 10 mg L−1. At the Zeekoegat WWTP, as well, 50% of the collected water samples exceeded the 1 mg L−1 phosphate limit (Hohls et al. 1998). Whereas the phosphate standard showed only limited success, it remains in force in these catchments, and was not adjusted there or applied to other water reservoirs.

Low- or Zero-Phosphate Detergents

Phosphate detergents have been shown to contribute 40% of the total phosphorus in wastewater (Pillay 2001). In SA, 32% of phosphate in domestic wastewater comes from detergents and these, on average, contribute 30% of the nutrient load to receiving waterbodies. Therefore, the use of low- or zero-phosphate detergents was evaluated as a possible strategy for reduction of external nutrient loading into water impoundments. In a study commissioned by the Water Research Commission (WRC), it was reported that depending on the catchment demographics, the use of low- or zero- phosphate detergents could achieve phosphorus reductions between 3 and 35%, with an average of 12% in water impoundments (Quayle et al. 2010). It was established by extension, that the introduction of low- or zero-phosphate detergents would reduce the costs of potable water treatment. For example, at the Hartbeespoort, Klipvoor and Roodeplaat impoundments, the estimated total savings that would be achieved by using low- or zero-phosphate detergents amounted to R616,134 (US $32,000) per year (Graham et al. 2012). Challenges to implementation of low- or zero-phosphate detergent regulations emanated from the associated costs. There were legal questions about whether consumers or detergent producers should bear the costs of producing low- or zero-phosphate detergents. The largest local producer of detergents in SA; however, voluntarily reduced phosphate concentrations in their products. Recently, it was reported that there were reductions in the phosphate loads in some river systems around SA which were speculated to have resulted from the reduced phosphate concentrations in detergent products.

Overall measures to prevent eutrophication in SA have focused on phosphorus management. Generally, these have produced minimal positive outcomes as eutrophication remains prevalent in SA impoundments. Therefore, prevention of eutrophication in SA impoundments remains imperative. Currently there are calls for the government to revisit the phosphate standard, modify it, and apply it in all WWTPs. Technically this calls for an overhaul of the wastewater treatment services to increase their capacity and their efficiency of nutrient removal. Implementation of the low- or zero-phosphate detergents also remains a viable option that can be explored for effective eutrophication control in SA. It remains relevant to ease the pressure on poorly performing WWTPs and to minimize nutrient pollution from communities that clean their laundry in waterbodies (Botha 2015). On the other hand, there is need to map out and identify the key/primary drivers of eutrophication for each water body.

Within-Waterbody Eutrophication-Remediation Measures in Selected Water Bodies

Despite the implementation of the phosphate standard, water quality and trophic status of water impoundments deteriorated, necessitating within-waterbody remediation to ameliorate the symptoms of eutrophication. Some of the within-waterbody remedial techniques are summarized in Table 2 and those implemented in some SA water impoundments are further discussed in subsequent sections.

Table 2 A summary of in-waterbody eutrophication remediation techniques (Hobson et al. 2012; Pandhal et al. 2018; Ugochukwu and Nukpezah 2008; Zamparas and Zacharias 2014)
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Epilimnion Mixing at Rietvlei Impoundment

Rietvlei impoundment is a small (2.06 km2) man-made waterbody in Gauteng Province, with a mean depth of 6 m and a volume of 12 × 106 m3 (Barnard et al. 2014). The waterbody is economically relevant as a source of raw potable water for the Rietvlei WTP and a destination for recreational water activities. The Hennops River, which is the only tributary for the impoundment, drains a catchment area consisting of two WWTP’s (Esther Park and Hartbeesfontein), human settlements, agricultural land, urban developments, and a nature reserve. Land use within the catchment area is limited to agricultural activities, mostly cattle rearing and maize farming. Therefore, possible sources of nutrients to Rietvlei are the WWTP’s effluents, agriculture runoff and animal waste from the nature reserve. Although there is a wetland upstream of the dam, which probably filters nutrients from the inflow, it cannot be ascertained if it still serves that purpose (Rossouw et al. 2008). However, as per the GD report of 2022, phosphate concentrations in Hartbeesfontein and Esther Park effluents were 0.33 and 0.96 mg L−1, respectively, which are below the stipulated 1 mg L−1 (Department of Water and Sanitation 2022c). Therefore, it is safe to assume that the contribution of effluent from WWTP’s to nutrient loading at Rietvlei is lower than in most impoundments in SA.

Even though nutrient loading patterns at Rietvlei have not been fully ascertained or clearly mapped, the trophic status of the waterbody is evidence that the nutrient inflow to the waterbody exceeds its eutrophication threshold. In a survey conducted by van Ginkel et al. (2004), it ranked as the most eutrophic impoundment amongst those studied. At the time, characteristic phytoplankton blooms were dominated by the dinoflagellate Ceratium hirundinella (van Ginkel 2004). However, as the water quality deteriorated, the phytoplankton community evolved and became dominated by cyanobacteria, particularly Microcystis aeruginosa. It was after the prevalence of cyanobacterial blooms in the impoundment that the nearby Rietvlei WTP started to experience challenges in its production, triggering the need for within-waterbody rehabilitation (van Vuuren 2012).

Based on impoundment morphometry, water uses and nutrient inventories, epilimnion mixing, using SolarBee© devices, was implemented as the most feasible and effective strategy (van Vuuren 2012). SolarBee© devices are solar-powered, long-distance reservoir circulators used for epilimnion mixing in the control of eutrophication. Epilimnion mixing ameliorates the symptoms of eutrophication through disturbance of the epilimnion and maintenance of conditions that suppress cyanobacterial growth but favor algae and complex aquatic food webs. The mechanism by which cyanobacterial growth is suppressed is not fully understood and several hypotheses have been proposed. The most plausible hypothesis is that the circulation interferes with cyanobacteria buoyancy regulation, thus they fail to migrate between the bottom of the water column where they absorb nutrients and the epilimnion, where they trap light energy for photosynthesis (Fan et al. 2014). SolarBee© devices are advantageous over other epilimnion circulators or mixers because they use solar energy, which minimizes operational costs. At Rietvlei, 16 SolarBee© units were installed across the whole impoundment between 2008 and 2009 (Hart 2012).

Immediately after the installation of the SolarBee© devices, there was an improvement in water quality, evident by the reduction of turbidity. Phosphate and nitrate concentrations in the water declined, despite the external nutrient loads remaining unchanged. Furthermore, the growth of cyanobacteria was suppressed in favor of green algae (van Vuuren 2012). However, decreases in nutrient concentrations were not sufficient to improve the trophic status of the waterbody, as phosphate, nitrate and chlorophyll-a levels remained high, in the eutrophic range (Barnard et al. 2014). The benefits of improved water quality, extended to Rietvlei WTP, as the cost of potable water production declined during the period of clear water (Booyens et al. 2012). However, a study conducted between 2010 and 2012 indicated that improvements in water quality were short-lived. During that period, phosphate concentrations increased steadily, from 0.13 to 0.40 mg L−1, and chlorophyll-a concentrations indicated a eutrophic to hypereutrophic state. Cyanobacterial blooms occurred in the wet season and were dominated by Microcystis sp. and microcystin concentrations in the impoundment were > 1.5 μg L−1 (Mbiza 2014).

The use of SolarBee© devices to control eutrophication has produced both positive and negative results. For example, at Crystal Lake, East Gravel Lake 4, and Lake Palmdale, USA, solar-powered epilimnion mixers were used successfully to control cyanobacterial blooms. In all three waterbodies there was a significant reduction in cyanobacterial cell density by ~ 82–95%, and green algae and diatoms became the dominant phytoplankton populations during the first and second year of application (Hudnell et al. 2010). Nonetheless, the effectiveness of circulation devices in eutrophication control relies heavily on continuous operation, as intermittent circulation may result in cyanobacterial growth. It is also crucial that the installed devices cover the whole waterbody to ensure adequate circulation. These therefore necessitate knowledge of the lake’s morphometric, physical, chemical, and biological characteristics to enable optimization of device locations, spacing, circulation rate and area, and intake depth (Hobson et al. 2012). At Rietvlei, the 16 devices were strategically positioned to cover the whole waterbody. However, a major drawback with Solarbee© application in SA is that the devices must be serviced by technicians based in the USA, which extends the time it takes to fix a broken one. When one or more devices do not work, cyanobacteria blooms can occur because physico-chemical conditions in the water remain favorable for cyanobacterial growth.

Circulation devices are strictly a curative measure, used to ameliorate the symptoms of eutrophication, without significantly reducing nutrient concentrations in the waterbody. Ultimately, to achieve long-term beneficial effects, external nutrient loads entering the waterbody must be reduced. For the Rietvlei water impoundment, it is crucial to identify and quantify nutrient loads from external point and non-point sources, as well as internal sources. Subsequently, an integrated rehabilitation program can be designed to target all the sources of nutrients at Rietvlei. For this water impoundment, animal waste is of particular interest as it possibly contributes significantly to nutrient loading. This is because animal waste was responsible for the eutrophication of the Nhlangazwane reservoir and subsequently the death of wildlife from cyanotoxin poisoning (Buss and Bengis 2012). Rehabilitation and maintenance of the upstream wetland is also crucial to filter nutrients from the water impoundment inflows.

Algaecide Application at Roodeplaat Impoundment

The Roodeplaat impoundment is a small (4.39 km2) man-made reservoir with a mean depth of 10.6 m and maximum volume of 41 × 106 m3. It is an economically important waterbody that is used for irrigation, potable water supply and recreational and sporting activities such as fishing, swimming, boating, jet-skiing, rowing, and canoeing. It is the only locally accredited site for local and Olympic rowing. Within the impoundment’s vicinity, several ecotourism activities take place, including camping, picnic spots, lodging, and meeting facilities (Silberbauer and Esterhuyse 2014). Roodeplaat receives its inflow from three main tributaries, the Pienaars, Hartbeesspruit and Edendalespruit Rivers. These rivers drain a catchment area of about 684 km2, characterized by formal and informal human settlements, agricultural activities, and urbanization. Approximately 50% of the nutrient load at Roodeplaat emanates from non-point sources including informal settlements and agricultural runoff. Two WWTPs, Baviaanspoort and Zeekoegat, discharge effluent into the rivers that feed into the reservoir, accounting for the remaining 50% of the nutrient load to Roodeplaat (Zulu 2020). According to the 2022 GD report, the Zeekoegat and Baviaanspoort plants obtained GD scores of 61% and 57%, respectively, indicating that both plants discharge poorly treated effluent into the catchment and ultimately into Roodeplaat reservoir. The Baviaanspoort plant is the most polluting, as it operates at 153% of its design capacity and releases partially treated or raw sewage into the Piennars River (Department of Water and Sanitation 2022c). On the other hand, effluent from the Zeekoegat WWTP passes through an artificial wetland before reaching the impoundment which filters nutrients from the effluent (Venter 2013). Both plants have been the subject of investigation by the South African Human Rights Commission over their poor performance (South African Human Rights Commission 2021).

Characteristic of waterbodies that receive nutrient-rich inflows, Roodeplaat has been classified as hypereutrophic since 1998 (van Ginkel 2004). In the period 2001–2021, it experienced a steady increase in concentrations of phosphates, nitrates, and chlorophyll-a, with the trophic status alternating between eutrophic (2001–2006 and 2009–2014) and hypereutrophic (2007–2008 and 2015–2021) (Mnyango et al. 2022). Water quality problems linked to high nutrient loads at Roodeplaat include protracted periods of cyanobacterial blooms during the warm summer months (November, May, or June) (Conradie and Barnard 2012). Lately, hyacinths have also become prevalent and grow massively, sometimes covering the entire water surface (van Ginkel and Silberbauer 2007).

Deteriorated water quality in Roodeplaat severely compromises the uses of the waterbody for recreational and sporting activities, and as a source of potable water, thus necessitating urgent rehabilitation. In 2020, ahead of the World Rowing Masters Regatta competition, the organizing committee commissioned Blue Green Water Technologies to treat the impoundment to mitigate a severe cyanobacteria bloom (Microcystis spp). The company uses a chemical called Lake Guard Oxy™, which is a buoyant, encapsulated form of sodium per-carbonate (SPC) (Sukenik and Kaplan 2021). Upon application onto the water surface, hydrogen peroxide is released, and it generates reactive oxygen species, super oxide carbonate, and hydroxyl radicals to kill cyanobacterial cells. The defining feature of Lake Guard Oxy™ is the slow release of hydrogen peroxide, in exceedingly small concentrations that are non-lethal to non-target species. Use of an encapsulated form reduces the risks of handling high concentrations of hydrogen peroxide (Hobson et al. 2012).

After application of Lake Guard Oxy™ at Roodeplaat, cyanobacterial biomass was reduced by almost 99.99% within a few days, and as of January 2021, the water was still clear (Sukenik and Kaplan 2021). As expected, the algaecide killed cyanobacteria without shifting the water quality to a state unfavorable for cyanobacterial blooms; hence, cyanobacterial blooms and hyacinths reappeared in the impoundment. A similar scenario was observed in a waterbody in The Netherlands, where hydrogen peroxide was used to control a Planktothrix bloom. Although cells were reduced by as much as 99% in 10 days, the bloom reappeared several weeks later (Jančula and MarŠálek 2011). Indeed, it has been established that for long-term effects, algaecides, particularly hydrogen peroxide, should be applied repeatedly.

As of 2021, physico-chemical variables (nitrates, phosphates) measured at Roodeplaat were indicative of a hypereutrophic waterbody. The calculated average Water Quality Index (WQI) was 93.94, which is class 4 and indicative of very poor water quality (Mnyango et al. 2022). This WQI value means that the impoundment water is fit only for irrigation purposes and cannot be used for domestic, industrial, or recreational purposes without prior treatment. At present Roodeplaat impoundment still needs rehabilitation to improve its trophic status and restore its recreational and amenity value. The first step would be to rehabilitate and upgrade the two WWTP’s within the catchment. As recommended, the operation capacity of both plants must be increased, and advanced wastewater treatment technologies with high nutrient-removal efficiency must be incorporated (Zulu 2020). Because the waterbody has been receiving high nutrient loads, it is likely that internal nutrient concentrations are elevated. These also need to be removed, through chemical or mechanical within-waterbody rehabilitation strategies.

The Integrated Biological Remediation Programme at Hartbeespoort Dam (HBPD)

HBPD is a large (20 km2), man-made reservoir in the Northwest Province, with a mean depth of 9.6 m (van Ginkel and Silberbauer 2007). The water in this impoundment is used for commercial irrigation, potable water supply, aquaculture, and recreational activities. It also supports the property market, with high-capital investments in its vicinity (Venter 2012). The main tributaries to the HBPD, the Crocodile River (90% inflow) and Magalies River (10% inflow), drain a 4112 km2 catchment that has 10 WWTPs, commercial areas, agricultural activities, industries, as well as formal and informal human settlements (Harding 2004a). Recent nutrient loading studies at HBPD showed that 53% of the nutrients come from non-point sources, including agricultural runoff, untreated sewage from leaking and overflowing sewer systems, and informal settlements (Botha 2015; Dennis and Dennis 2019). The 10 WWTPs in the catchment are the major contributors of point source nutrient loads, as they discharge 650 megaliters of effluent per day into rivers that flow into HBPD (Cukic et al. 2012; Petterson 2019b). Some 60% of the WWTPs discharge poorly treated effluent that is noncompliant with the 1 mg L−1 phosphate standard. In the 2022 GD report, only one plant (Esther Park WWTP) with a daily discharge of 0.74 megalitres, obtained the GD certificate (95% score).

Because of the high nutrient inputs, HBPD is renowned globally as the most hypereutrophic reservoir and its water quality has progressively deteriorated over the years (Oberholster and Botha 2010; van Ginkel 2004). The effects of eutrophication in HBPD manifest as massive, seasonal cyanobacteria blooms. Typically, these blooms appear as hyperscums (floating mats) of the buoyant cells of M. aeruginosa that accumulate along the dam wall during calm weather conditions (Oberholster and Botha 2010). Another characteristic symptom of eutrophication in the reservoir is the loss of ecological integrity, as characterized by distorted food webs and reversed energy flow. Because of the destruction of littoral vegetation, populations of macroinvertebrates and zooplankton declined, having lost their habitats and shelter from predators. Physico-chemical properties in the water favor the growth of less palatable algae, which reduces zooplankton predation. Therefore, zooplankton populations were reduced and the fish that prey on zooplankton also declined. Consequently, undesirable species of fish like Clarias gariepinus (catfish), Cyprinus carpio (common carp), and Chetia flaviventris (canary kurper) became dominant (Amorim and do Moura 2021). Although these fish are omnivorous, they are undesirable for a healthy aquatic food web because they have negative impacts on zooplankton. Catfish larvae feed on zooplankton and the adult carp prey on the eggs and young of desirable fish species. Both species feed on zoobenthos and in the process uproot aquatic vegetation, thereby disturbing bottom sediments and releasing nutrients into the water column (Havens 2008; Qin et al. 2013).

The Hartbeespoort Dam Integrated Biological Remediation Programme, locally referred to as Harties Metsi a Me (HMaM) was initiated in 2005 for the prevention, control, and remediation of eutrophication (Fig. 4) (Harding 2004a, b). The proposed first phase of the program encompassed three within-waterbody remediation techniques (sediment dredging, biomass harvesting, and floating wetlands), which were all intended to reduce internal nutrient loads and limit nutrient supply to Microcystis sp., thus restricting its growth (Harding 2004a, b). The proposed second phase was designed to reduce external nutrient loads from both point and non-point sources within the catchment. This was to be achieved through the rehabilitation of wetlands, improved compliance with discharge regulations at point sources, as well as nutrient removal by chemical applications (Venter 2012). The first phase of the program was implemented partially in 2005 with physical removal of algae and hyacinths, establishment of floating wetlands, and fish harvesting.

Fig. 4
figure 4

Schematic presentation of the Hartbeespoort Integrated Biological Remediation Program

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Physical Removal of Algae Biomass and Hyacinths

The decay of aquatic biomass contributes to internal nutrient loads in water impoundments, which in some cases may account for > 50% of the total nutrient load (Welch and Cooke 2005). Therefore, removal of living biomass (aquatic plants or algae) is an effective approach to curb further growth of biomass and improve water aesthetics, water quality and ecosystem functions (Croucamp and Venter 2012). Such action may also reduce internal nutrient loading by removal of nutrients trapped in biomass, thereby preventing biomass accumulation in sediments and its subsequent decomposition and nutrient release (Pandhal et al. 2018). Harvesting a hectare of water hyacinths from a waterbody translates to removal of 1.8 and 0.8 tons of nitrogen and phosphorus, respectively (Petterson 2019b). However, in some cases, removal of biomass opens space for the growth of new cells (Hao et al. 2016). At HBPD, physical removal of algal biomass and hyacinths was incorporated into HMaM. By 2012, an estimated 16 million liters of ‘algal soup’ and 44,280 m3 of hyacinth were harvested at a cost of R1 million (US $52,000) per ton. These were converted into approximately 280 tons of compost and vermicast in a vermiculture beneficiation project (Croucamp and Venter 2012). Immediate effects of biomass harvesting at HBPB included an improvement in the impoundment aesthetics and to a limited extent, the water quality. Removal of the algal biomass reduced the turbidity, and the water became clearer. Removal of the water hyacinths opened space for water-based activities in the impoundment (van Ginkel et al. 2002).

Floating Wetlands

At HBPD, more than 300 floating wetland units (16 m2 each), covering an area of approximately 5000 m2, were commissioned on selected sites along the shoreline. Floating wetlands are an ecotechnology in which emergent plants are grown hydroponically on floating supports with their roots immersed in the water, where they assimilate nutrients directly from the water column and entrap suspended particulates (Yeh et al. 2015). Submerged parts of floating wetlands provide substrate for microbial activity, which contributes to nutrient cycling, and provides additional food for macroinvertebrates and fish (Vymazal 2007). These floating structures support a diverse community of macrophytes, macroinvertebrates, zooplankton, fish, and aquatic birds. Macrophytes are important primary producers, and they absorb nutrients from the water column, thereby reducing the amount of nutrients available for undesirable cyanobacteria. Furthermore, they act as shelters, and roosting and nesting spots for aquatic birds that feed on fish. Zooplankton hide and breed among the submerged roots and feed on cyanobacteria, preventing their overgrowth into blooms. Overall floating wetlands serve the functions of nutrient uptake and restoration of distorted food webs.

At sites where floating wetlands were established, there were observable improvements in water quality and aquatic ecosystem ecological functioning. Water quality improvements included reductions in turbidity, soluble reactive phosphorus, and chlorophyll-a concentrations, with the water remaining clear (van Ginkel et al. 2012). Ecological improvements included the appearance of macrophytes and filamentous algae and increases in zooplankton species richness and diversity on sites where floating wetlands were installed. Subsequently the growth of cyanobacteria was suppressed, and the water remained clear (van Ginkel et al. 2012). However, this healthy ecosystem state did not last, as dense algal blooms occurred in the subsequent wet season, resulting in a macrophyte die-off. The effectiveness of floating wetlands in ameliorating eutrophication and cyanobacterial blooms is limited as it depends on the ability of the higher plants to successfully compete with bloom-forming cyanobacteria for nutrients and sunlight. By their nature, bloom-forming cyanobacteria, are typically highly competitive, and although they may be outcompeted in the short term, they eventually re-establish and proliferate into blooms (Carey et al. 2012).

Fish Harvesting

Biomanipulation is the restructuring of food webs and is sometimes used as a remediation tool to shift eutrophic waters from a turbid state dominated by phytoplankton to a clearer state, dominated by aquatic macrophytes (Jeppesen et al. 2012). In a ‘top-down’ approach, grazing pressure on cyanobacteria is increased by enhancing predatory interactions at higher trophic levels. In the ‘bottom-up’ approach, however, efforts are made to minimize nutrient availability for cyanobacteria productivity (Sierp et al. 2009). Successful biomanipulation in aquatic ecosystems should result in lower turbidity, and reduced concentrations of total phosphorus, chlorophyll-a, and algal biomass (Peretyatko et al. 2009).

At HBPD, a ‘top-down’ approach was implemented during which individuals from populations of the three putatively zooplanktivorous ‘coarse’ fish were removed using large-catch techniques. The initial target was to remove 200–300 tons of fish in the first year and achieve approximately 20% stock reductions in the second year. The desired effects of fish harvesting were a consumer-mediated trophic cascade whereby zooplankton populations would increase, and phytoplankton populations would decrease (Ekvall et al. 2014). It was hoped that this would cause the reemergence of desirable omnivorous, algae-eating, palatable fish species with high commercial value. Furthermore, removal of Clarias and Cyprinus spp. was expected to reduce nutrient recycling from the sediments (internal loading), caused by fish foraging activity (Venter 2012).

Over the program's duration, 177 tons of fish were removed at HBPD, less than the initial target. Therefore, the effects of fish removal were minimal, with only the biomass of Mozambican tilapia increasing, while increases in other desirable fish species were minimal (Hart and Matthews 2018). The utility of biomanipulation as an ecological approach to rehabilitate reservoirs is disputed. The few reports of their successful implementation come from application in shallow waterbodies, but there has been no confirmation regarding their long-term success (Kasprzak et al. 2003). Furthermore, the fundamental prerequisite for a successful biomanipulation program is a thorough understanding of and proper manipulation of trophic interactions (Ugochukwu and Nukpezah 2008). The trophic interactions at HBPD negated any positive outcomes of the implemented fish removal exercise. In their study, Harding and Hart (2013) noted that the three fish species removed from HBPD were principally benthic feeders, hence their removal had minimal effects on zooplankton populations. The foraging activity of carp and catfish was found to contribute minimally to sediment nutrient recycling, hence their removal had insignificant effects on internal nutrient reduction (Hart and Harding 2015).

Effectiveness of Rehabilitation at HBPD

The major goals of the HMaM rehabilitation program were to restore its ecological functioning, improve water quality and usability, as well as reduce water-column phosphate concentrations from 0.2 to 0.05 mg L−1 (Harding 2004a). Overall, the success of HMaM was minimal and limited to local, short-term improvements in water quality, ecological function, and aesthetics within the first 18 months of implementation. Whereas these were important in restoring the recreational uses of the waterbody, they were not enough to lower the trophic status and restore ecological integrity. Also, there were uncertainties as to whether the improvement in water quality could be directly attributed to HMaM interventions (Hart and Matthews 2018). For example, reductions in chlorophyll-a concentrations recorded in 2007 and 2008 at HBPD, were also observed in control sites where remedial measures were not implemented. Also, in situ and remote satellite imaging water quality data collected between 1980 and 2020 did not reveal improvements in water quality and productivity attributable to HMaM interventions (Ali et al. 2022). Furthermore, measurements of pH, temperature, and phosphate concentrations at HPBD in the period between 2010 and 2012, showed no improvements in the water physico-chemical parameters as they remained favorable for the growth and proliferation of cyanobacteria (Mbiza 2014). With respect to the internal nutrient loads at HBPD, the implemented within-waterbody remediation techniques failed to reduce internal nutrient loads, as nutrient loading modelled between 2010 and 2017 showed a steady increase in nutrient inputs into HBPD and a reduced ability of the waterbody to release nutrients via the outflow. An estimated 39% of TN and 53% of TP inputs were released through outflows, which means that 61% of TN and 47% of TP inputs remained in the dam (Carroll and Curtis 2021). This was confirmed by Botha et al. (2015) who also reported that only 30% of phosphate inputs were released via the outflows and a staggering 70% remained in the sediment (Botha 2015). This implies that HMaM interventions like fish, algae and hyacinth harvesting made an insignificant contribution to nutrient removal since the rate at which nutrients were removed by those efforts were much lower than the rate at which total phosphorus accumulated in the aquatic system (Carroll and Curtis 2021).

The key limitation of the strategies implemented at HBPD is that they were curative means to treat the symptoms of eutrophication, with a limited ability to reduce the nutrient inventories. Indeed, there is substantial evidence that within-waterbody remedial techniques such as biomass harvesting, sediment dredging, and aeration are ineffective in reducing nutrient concentrations (Kibuye et al. 2020; Schindler 2006). Furthermore, the short duration of the HMaM contributed to its limited success because it takes a long time to remediate eutrophic water bodies to their former oligotrophic or mesotrophic state (Chorus et al. 2020).

Future Rehabilitation Strategies at HBPD

Effective rehabilitation of the HBPD requires that the phosphate concentrations be reduced to as low as 0.05 mg L−1 and to reach that threshold, nutrient load reductions of > 75% are required (Harding 2008). This could be achieved through effective removal of nutrients from WWTPs and those from non-point sources such as run-off from informal settlements and agriculture. To address the issue of high nutrient loads from point sources, stricter effluent discharge standards can be enforced. However, the DWS acknowledges the challenges associated with the enforcement of national legislation protecting water resources. Non-point sources as well are remarkably diverse and difficult to manage. Although agricultural runoff can be minimized through best agricultural practices, fertilizer use in SA is not adequately quantified and the extent of nutrient loading from agriculture is not documented. Therefore, it may be challenging to implement strategies to minimize agricultural runoff (Van Der Laan and Franke 2019). Nonetheless, the rehabilitation of wetlands, and the restoration of the shoreline can be effective in filtering nutrients from nonpoint sources. Alternatively, a revised strategy that will incorporate the treatment of incoming water before entering the dam in addition to the measures of treatment taking place in WWTPs and within the waterbody can be implemented (Cukic et al. 2012; Petterson 2019a).

Preimpoundment stands out as the most feasible approach for the treatment of incoming water before entering the impoundment. In preimpoundment, inflows are diverted, impounded, and treated before entering the reservoir. Pre-impoundment has been used successfully for sediment settling in irrigation projects and is recommended as an effective eutrophication control strategy if combined with chemical nutrient removal (Douglas et al. 2016; Ugochukwu and Nukpezah 2008). The feasibility of this concept was previously investigated at HBPD, and it was concluded that nutrient load reductions from the inflow at the point of entry into the reservoir were possible but required further scientific, demographic, and legal evaluations prior to implementation (Twinch and Grobler 1986). Furthermore, implementation requires careful selection of a chemical that can be used to precipitate phosphorus and the preferred option is one that is easy to apply, low-cost, not lethal to aquatic life and does not cause drastic and detrimental changes to the physico-chemical properties of the water (Douglas et al. 2016). Currently, Phoslock®, acid mine drainage (AMD) and calcium hydroxide are some of the chemicals (materials) that have been evaluated for use in chemical nutrient removal from HBPD inflows (Atta et al. 2020; Mitchell et al. 2016; Ross and Cloete 2006). Of these three, acid mine drainage is more attractive. HBPD receives substantial amounts of AMD from the Western basin, which could be used to precipitate phosphate from inflows cost effectively (Hobbs 2017). The mine water from the Western basin can be piped to a floating phosphate removal plant at the three waterbody inlets (Magalies, Crocodile and Swart Rivers), and aerated to precipitate phosphate as FePO4 and Fe2(PO4)3 (Mitchell et al. 2016). Previous studies showed that the infiltration of additional mine water to the impoundment would have no significant effect on limnological conditions, except for increasing sulfate concentrations. AMD has been used successfully as a coagulant to remove phosphorus from municipal wastewater, with the removal efficiency dependent on the dosage of AMD and the initial water pH (Ruihua et al. 2011; Smyntek et al. 2022). Chemical phosphate removal processes generate phosphate-rich sludge as waste, necessitating its removal through sediment dredging. The sediment dredging operation can be extended to the dam basin to reduce internal nutrient loads, minimize nutrient recycling from sediments as well as increase dam volume. Because of the large size of the reservoir, sediment dredging would be prohibitively expensive, hence a cost-effective dredging system was designed (Cukic and Venter 2010). To help recoup dredging costs, the nutrient-rich sediment can be used in agriculture to improve sandy soils (Kazberuk et al. 2021). After separation, the solid phase can be used to manufacture construction products such as sand, bricks, blocks, and aggregates (Crocetti et al. 2022). Phosphate from the sediment can also be recovered and used to produce phosphoric acid (Masindi and Foteinis 2021).

Conclusions

Eutrophication of surface water in South Africa is an old problem that still threatens water quality. Most river systems and artificial impoundments are eutrophic to hypereutrophic. The socioeconomic impacts of eutrophication are felt in all aspects of the economy and there is evidence that these have not been completely quantified. Therefore, it is not possible to fully appreciate the impact of eutrophication on economic sustainability and development. While groundbreaking research from South Africa was significant in raising a global awareness of this problem, and the country was amongst the first to promulgate standards aimed at preventing eutrophication, the country has not been able to control eutrophication effectively and pragmatically. Attempts to ameliorate eutrophication have mostly been curative and have fallen short in addressing how to reduce nutrient loading into water bodies. These curative efforts have produced short-term beneficial effects on water quality and aesthetics but have contributed only minimally to reducing the trophic status of water impoundments, while incurring prohibitive costs. Moving forward, it is crucial to shift away from treating the symptoms of eutrophication and focus on its prevention. For example, most water treatment plants within the catchment area of eutrophic waterbodies have undergone upgrades to increase their efficiency in treating eutrophic waters to acceptable standards, whereas wastewater treatment plants have not received similar attention. Tailor-made, integrated management initiatives that address issues of nutrient loading from point sources and diffuse sources in the watershed remain key in addressing the issue of eutrophication.