Unveiling the Hydrochemical and Ecotoxicological Insights of Copper and Zinc: Impacts, Mechanisms, and Effective Remediation Approaches

Abstract

Water pollution is a pressing global issue significantly affecting ecosystem health, biodiversity, and human well-being. While numerous studies have concentrated on toxic metals like cadmium, lead, and mercury, essential metals such as copper and zinc often receive less attention. This review focuses on the distribution and occurrence of copper and zinc in surface water, their accumulation in freshwater organisms, and potential strategies for mitigating the environmental pressure caused by these metals. Zinc concentrations in uncontaminated freshwater usually range from 3 to 12 μg∙L⁻¹ and form low-bioavailable hydroxo-complexes that are especially stable in weak alkaline water. The zinc concentration trend globally is Europe > Africa > Asia > South America > North America. Conversely, copper concentrations vary from 0.2 to 5.5 µg∙L⁻¹, with the order being Asia > Africa > South America > North America > Europe. Humic substances are the likely predominant ligands for copper in these environments. The accumulation of copper and especially zinc in freshwater animals may not be a reliable indicator of metal pollution due to potential metabolic regulation. Bioremediation approaches, including phytoremediation and biosorption using plants and microorganisms, show promise in addressing water contamination. Future research should emphasize advanced bioremediation methods, emission reduction strategies, and refined modeling techniques to predict pollution trends and evaluate remediation effectiveness.

1. Introduction

Water pollution is a pressing global issue with far-reaching consequences for ecosystem health, biodiversity, and human well-being. In developing countries, 70% of industrial waste goes untreated into water, while 1.8 billion people globally use contaminated water sources for their daily needs. Water pollution causes economic damages of up to $4.8 billion per year in Middle Eastern countries [1]. The World Bank estimates that the global economic cost of water insecurity reaches $500 billion annually [2]. Coming from various sources, including industrialization, agricultural activities, natural factors, and inadequate water supply and sewage treatment facilities, it poses significant threats [3]. Industrial activities, such as those in heavy industry, textile industries, nanotechnology, and rapidly developed solar photovoltaic and wind turbines manufacturing, release a mixture of toxic chemicals, among them cadmium, zinc, lead, copper, and chromium, into water bodies, contributing substantially to pollution. Likewise, agricultural practices involving pesticides, fertilizers, and organic wastes contaminate water with harmful substances like aromatic compounds and their metabolites, chloro-derivates, nitrogen-containing heterocycles, metal-containing compounds, posing risks to both environmental and human health [4,5,6]. Furthermore, natural factors, such as weathering and geological processes, can affect the physico-chemical profile of pollutants and their mobility, amplifying their harmful effects on both wildlife and humans, thereby exacerbating the issue [7]. This scenario leads to heightened exposure to industrial chemicals and pathogens [3] on one hand, while on the other hand, it underscores the critical need for monitoring and analyzing the chemical profile of water bodies.

The chemical profile of water serves as a valuable indicator of its overall quality. Recent studies have revealed critical statistics, indicating that approximately 96%, 63%, and 74% of river samples in India surpass the WHO recommendations for boron, barium, and aluminum concentrations, respectively [8]. Through the analysis of hydrochemistry and standardized water quality parameters, including metal concentrations, NO₂⁻, Cl⁻, and SO₄²⁻, coupled with innovative spatio-temporal modelling techniques as an example [9], researchers can uncover the complex interaction mechanisms between water and its surrounding environment. This holistic approach not only provides insights into the dynamics of chemical pollution and regional water quality status but also offers valuable guidance for policymakers at both national and international levels, facilitating informed decisions regarding resource management [9,10]. Moreover, various indices such as sodium adsorption ratio, permeability index, water quality index, and health risk assessment play crucial roles in evaluating water suitability for various purposes, ranging from agricultural irrigation to human consumption [11,12].

Numerous narrative reviews have extensively documented the hydrochemistry and ecotoxicological impact of transition metals. While many studies have focused on technogenic toxic metals such as cadmium, lead, and mercury [13,14,15,16], essential metals like copper and zinc have often been overlooked. Nevertheless, projections suggest that by 2050, alongside the production peaks of cadmium, lead, and chromium [17], attention may shift towards copper and zinc [18], potentially altering the emission dynamics of metals in nature. This study aims to (1) analyze the distribution and occurrence of copper and zinc in surface water; (2) explore potential correlations in the absorption capacity of these metals; (3) assess the peculiarities of Zn and Cu accumulation in freshwater animals, primarily fish, abundant groups within freshwater ecosystems; and (4) outline promising approaches for mitigating the environmental pressure exerted by these metals.

2. Zinc

2.1. Occurrence of Zinc in Surface Waters and Sediments

Zinc, ranked as the 24th most abundant element [19], is released into the environment through both natural processes and human activities, with human-induced sources surpassing natural ones. Mining, metallurgy, and zinc-containing products significantly contribute to its dispersion in air, water, and soil [20,21]. Additionally, runoff from municipal facilities, renewable energy installations, and agricultural practices also contribute to zinc influx [22,23]. The global production of zinc is expected to grow by 2% to 12.5 Mt in 2024 [24], resulting in a surplus. While zinc tends to remain bound to soil and sediment in water bodies, leaching may occur at waste disposal sites. For instance, in Japan alone, approximately 641 tons of water-soluble zinc-containing compounds and, in Germany, 3.187 tons of Zn were discharged into water bodies in 2018 and 2000, correspondingly [20,23]. According to the Pollutant Release and Transfer Register, zinc ranked as the third most-released chemical into water bodies in that time [20]. The fact that zinc was pointed out as a priority pollutant in the USA, called out as a potential specific pollutant in the UK, and categorized as a List II dangerous substance in the EU necessitates the establishment of pollution reduction programs. Similarly, the UK has identified zinc as a potential “specific pollutant” due to the significant quantities discharged into water systems [25]. Moreover, some findings have claimed a significant increase in Zn demand in the near future [26], which potentially goes hand-in-hand with an increasing risk of pollution.

In Germany, both industrial and urban activities have contributed to zinc contamination in rivers like the Rhine and the Ruhr river basin, which contribute to approximately 50% to total zinc load transported into the River Rhine [21]. As an example, in 2000, the Zn input from industrial discharges in the Rhine catchment area was found to be of 1.0 g km⁻² d⁻¹ [23], potentially leading to increased zinc contamination in surface water. Consequently, high concentrations of up to 150 µg L⁻¹ Zn have been detected in Ruhr, likely originating from geogenic backgrounds or past mining activities [27]. However, there has been a positive trend indicating a reduction in contributions from industrial and urban pollution over time [27], suggesting some success in pollution management efforts (Table 1).

Similar trends have been observed in other regions. In Saxony state, Germany, declines in zinc concentrations have been noted over the years, alongside reductions in industrial and urban pollution inputs. In particular, the average concentration of Zn in the suspended particulate matter in Elbe River was determined to be 676 mg·kg⁻¹, and it declined by 58.2% over the course of nearly 20 years in the early XXI century [28]. In Poland, the Krzna River has shown slight zinc contamination in sediments, with zinc concentrations ranging from 16.2 to 23.5 mg·kg⁻¹. However, these levels remain four times lower than national averages of 73 mg·kg⁻¹. Meanwhile, zinc was found to be the most abundant metal in the Krzna River, followed by lead, nickel, copper, and cadmium [29].

Table 1. Temporal and continental data on Zn and Cu occurrence (µg·L⁻¹) in surface waters and sediments [30].

	1970s	1980s	1990s	2000s	2010s
	Concentrations over time
	Mean ± S.E.
Zn	52.2 ± 24.4 (12)	60.2 ± 30.5 (10)	1948.7 ± 394.7 (13)	842.8 ± 649.1 (83)	1180.1 ± 456.9 (74)
Cu	5.9 ± 1.6 (12)	8.5 ± 6.1 (10)	84.6 ± 53.6 (6)	433.6 ± 307.7 (90)	120.0 ± 39.3 (61)
	Concentrations across five continents
	Africa	Asia	Europe	North America	South America
Zn	1169.0 ± 680.7 (42)	889.6 ± 48.9 (122)	1339.0 ± 979.9 (19)	86.9 ± 48.5 (7)	680.5 ± 608.4 (5)
Cu	190.8 ± 67.2 (37)	345.9 ± 246.4 (113)	14.6 ± 2.8 (18)	15.9 ± 9.8 (6)	142.6 ± 70.6 (6)

S.E.—standard error; values in brackets denote selected sample numbers.

Table 1. Temporal and continental data on Zn and Cu occurrence (µg·L⁻¹) in surface waters and sediments [30].

	1970s	1980s	1990s	2000s	2010s
	Concentrations over time
	Mean ± S.E.
Zn	52.2 ± 24.4 (12)	60.2 ± 30.5 (10)	1948.7 ± 394.7 (13)	842.8 ± 649.1 (83)	1180.1 ± 456.9 (74)
Cu	5.9 ± 1.6 (12)	8.5 ± 6.1 (10)	84.6 ± 53.6 (6)	433.6 ± 307.7 (90)	120.0 ± 39.3 (61)
	Concentrations across five continents
	Africa	Asia	Europe	North America	South America
Zn	1169.0 ± 680.7 (42)	889.6 ± 48.9 (122)	1339.0 ± 979.9 (19)	86.9 ± 48.5 (7)	680.5 ± 608.4 (5)
Cu	190.8 ± 67.2 (37)	345.9 ± 246.4 (113)	14.6 ± 2.8 (18)	15.9 ± 9.8 (6)	142.6 ± 70.6 (6)

S.E.—standard error; values in brackets denote selected sample numbers.

A particularly high concentration of zinc, 1.5 mg L⁻¹, was recorded in the Zletovska River, which is situated near the Pb/Zn mines in Zletovo, northeastern Macedonia. This area also exhibits significant co-pollution with cadmium (2.0 μg·L⁻¹) and manganese (2.5 mg L⁻¹) [31]. The zinc levels exceed Canadian and US EPA limits for aquatic organisms by 50 and 12.5 times, respectively [32], thereby potentially exerting adverse effects on biota and biodiversity.

On the other hand, the water reservoirs in West Pomerania (Poland), which are located near emission sources, exhibited an elevated average zinc concentration in the water, ranging from 26 to 63 μg·L⁻¹. The sediment’s zinc content, averaging from 25.8 to 118.2 mg·kg⁻¹ [33], exceeded national averages of 73 mg·kg⁻¹ [29]. Notably, the village pond contained the highest zinc content, similar to the findings that have been shown in our previous reports [34]. Particularly high zinc concentrations, reaching 314 μg·L⁻¹, were observed in the Dnister river basin located in rural West Ukraine, correlating with the intensity of agrotechnical activities. Conversely, the zinc levels in Kasperivtsi water reservoir, a part of National Nature Park “Dnister Canyon”, were below the detection limit [35].

In general, the concentration of zinc in freshwater bodies declined in the following order: Europe > Africa > Asia > South America > North America (Table 1,

Table 2. Total zinc concentrations (µg·L⁻¹) in global river and lake water bodies.

	Country	Place	Concentration Range: Surface Water (µg·L−1) Sediment (µg·kg−1 D.M.)	Ref.
Zn	ASIA
	India	South-eastern coastal: Andhra Pradesh	<0.070	Mahdi et al., 2022 [36]
	Thailand	Chao Phraya River (n = 16)	5.33–17.10 (rainy season); 2.40–19.50 (dry season)	Niampradit et al., 2024 [37]
		Chao Phraya River (n = 9)	28.06–160.60	Chanpiwat and Sthiannopkao 2013 [38]
		Tha Chin River (n = 38)	160–7470	Veschasit et al., 2012 [39]
	Cambodia	Tonle Sap—Bassac River (n = 11)	not detected	Veschasit et al., 2012 [39]
	Indonesia	Citarum River (n = 10)	3.28–44.26	Veschasit et al., 2012 [39]
		Winongo River (n = 8)	-	Fadlillah et al., 2023 [40]
	Malaysia	Linggi River (n = 15)	1.16–6.35	Razak et al., 2021 [41]
		Semenyih River (n = 8)	33.10–49.19	Al-Badaii and Shuhaimi-Othman 2014 [42]
	Vietnam	Saigon River (n = 8)	5.38–311.10	Chanpiwat and Sthiannopkao 2013 [38]
	Iran	Chah Nimeh reservoir—surface water (n = 7)	730–960 (in spring) 1390–1560 (in summer)	Bazrafshan et al., 2015 [43]
		Chah Nimeh reservoir—sediment (n = 7)	87,000–99,000 (spring) 88,000–99,000 (summer)
	China	Dianchi Lake: water surface/sediment	20.64/496,800	Liu et al., 2021 [44]
		Chaohu Lake: water surface/sediment	23.05/341,000	Wang et al., 2016 [45]
		Daye Lake: water surface/sediment	2.68–3.64/237,500	Wang et al., 2023 [46]
		Yangtze River: water surface/sediment	5.40/104,100	Li et al., 2020 [47]
	EUROPE
	Poland	Muchawka River (n = 16/12): surface water/sediment	15.3–20.1/16,200–21,300	Kluska and Jabłonska, 2023 [48]
		Liwiec River (n = 32) surface water/sediment	16.4–19.6/16,400–22,900
	Greece	Bay and Gulf of Thessaloniki (Aegean Sea including the Bay and Gulf of Thessaloniki)	16.5–75.9	Christophoridis et al., 2009 [49]
	Turkey	Ataturk lake–sediment	60,790	Karadede and Unlu, 2000 [50]
		The Dipsiz stream (sediment)—tributary of the river Buyuk Menderes	37.000 ± 26.000	Demirak et al., 2006 [51]
		The Dipsiz stream—tributary of the river Buyuk Menderes	1.051 ± 1.751
	Spain	Guadaira river: surface water/sediment	20–190/43,100–1,033,000	Enguix Gonzalez et al., 2000 [52]
		Tinto River (sediment)	110,000–6,730,000	Galan et al., 2003 [53]
	France	Cajarc site, Lot River (sediment)	909,000–10,000.000	Audry et al., 2004 [54]
	Germany	Malter Reservoir (sediment)	<1,900.000	Muller et al., 2000 [55]
	Scotland	Lochnagar (sediment)	39,000–180,000	Yang et al., 2002 [56]
	Switzerland	Lake Zurich (sediment)	50,000–675,000	Von Gunten et al., 1997 [57]
	Netherlands	Meuse River (sediment)	803,000–108,3000	Van der Berg et al., 1999 [58]
	AMERICA
	Mexico	Continental shelf—Tamaulipas (sediment)	17,300–115,700	Ponce 1995 [59]
		Continental shelf—Tabasco (sediment)	47,300–127,900
		Southeast Gulf of Mexico—Lagoons	0–85,200	Rosales et al., 1999 [60]
	United States	300 coastal and estuarine sites (sediment)	140,000 (mean)	O’Connor and Cantillo 1992 [61]
	U.S. Virgin Islands	Developed sites/undeveloped sites (wetland, forest, shrub, rangeland)	26,700–153,000/ 27,300–95,300	Lancellotti et al., 2023 [62]
	Urugway	Montevideo Harbor (sediment)	174,000–425,000 (summer) 183,000–491,000 (winter)	Muniz et al., 2004 [63]
	Brazil	Jurujuba Sound	15,000–337,000	Baptista et al., 2000 [64]
	Venezuela	Coral reef sediment	36,000–77,000	Bastidas et al., 1999 [65]
	Chile	Southern Fjords	91,000–122,000	Ahumada et al., 2015 [66]
	AFRICA
	Malawi	Lake Chilwa: surface water/sediment	6.24–1168.70/66,130	Mussa et al., 2019 [67]
	Cameroon	Municipal Lake (sediment)	26,800–341,000	Ekengele et al., 2008 [68]
	Egipt	El-Mex Bay	20,790–59,290	Abdallah, 2008 [69]
	Morocco	Sebou Estuary sediments	179,000	Cheggour et al., 2005 [70]
		Nador lagoon	4000–466,000	Bloundi et al., 2008 [71]
	Zambia	Kafue River—Copperbelt mining region	1000–125,000	von der Heyden and New 2004 [72]
	Namibia	sediments of the Gruben River	205,000	Taylor and Kesterton 2002 [73]
	Algeria	Tafna Wadi	14–70	Benmostefa et al., 2022 [74]
	Tanzania	Mwanza Gulf of Lake Victoria (sediment)	45,400	Kishe and Machiwa 2003 [75]

). In contrast, the concentration of copper followed a different pattern, declining in the order: Asia > Africa > South America > North America > Europe. Notably, copper concentrations never exceeded WHO and USEPA standards.

In European countries, environmental quality standards (EQS) have been established to regulate the total fraction of zinc in surface water. These standards vary depending on factors such as water hardness, with ranges set from 8 to 125 μg·L⁻¹. Specifically, in the UK and Wales, standards for dissolved bioavailable zinc have been set at 10.9 µg L⁻¹, while in Germany, where the ACR factor is equal to 3 and the average zinc concentration in freshwater reservoirs ranges from 3 to 12 μg·L⁻¹, the standard has been set at 33 μg·L⁻¹ [76].

The global scenario likely mirrors that of the EU (Table 2). For instance, in the most polluted downstream sections of the Umeda River (Japan) and the Nile River (Egypt), dissolved zinc concentrations ranged from 4.6 to 71.9 μg·L⁻¹ and 80 to 220 μg·L⁻¹, respectively [20,77]. However, in some cases, zinc pollution can reach significant levels, as observed in the low-river-flow conditions of the Chu-Talas Basin in Central Asia. Here, zinc concentrations ranged from 1104 to 9855 μg·L⁻¹, with an average value of 4972 μg·L⁻¹ [78]. Also, in Tai Lake, zinc concentrations were observed to range from 18 to 1246 μg·L⁻¹, with an estimated 50.7% of aquatic organisms potentially impacted by zinc [79]. Similarly, Liu et al., (2017) observed zinc concentrations in the Songhua River, China, varied from 920 μg·L⁻¹ to 70.81 mg·L⁻¹ [80].

In terms of particulate zinc, it was found in sediment within a wide range, from 28.59 to 2000 mg·kg⁻¹. Specifically, in the Umeda River, Yangtze River, and Nile River, particulate zinc levels varied from 0.42 to 2.01 mg·kg⁻¹ [20], 47.6 to 154 mg·kg⁻¹ dry weight [81], and 28.59 to 155.02 mg·kg⁻¹ [77], respectively. During weekdays, total zinc concentrations were notably higher, ranging from 15 to 43 μg·L⁻¹, compared to weekends where they ranged from undetected to 32 μg·L⁻¹, suggesting potential anthropogenic influences like industrial discharges [20]. However, overall enrichment factors for zinc generally remain below 2, indicating that zinc contamination might not be widespread in most studied regions [81]. Nevertheless, occasionally, the zinc levels in analyzed samples exceed guidelines for aquatic life, and Predicted No-Effect Concentrations for zinc vary from 24 to 840 μg·L⁻¹, depending on the species, necessitating strict monitoring due to expected increases in zinc application and emissions in surface and groundwater in the near future [24].

2.2. Zinc Mobility and Complexation

It is believed that in water bodies, Zn can exist in low-bioavailable hydroxo-complexes. The coordination of water to zinc exhibits a covalent character, in contrast to the electrostatic interactions observed with alkali metals like Na⁺ and K⁺. Zinc commonly coordinates with six water molecules to form the hexa-aquo complex [Zn(H₂O)₆]²⁺, but it can also form tetra-aquo [Zn(H₂O)₄]²⁺ and penta-aquo [Zn(H₂O)₅]²⁺ complexes. These complexes show flexible geometries due to the absence of ligand field stabilization in zinc’s d10 electron configuration. Zinc aquo complexes act as acids and undergo deprotonation, forming hydroxo complexes that impact the pH of solutions [82]. Hydroxo-complexes are particularly stable in the weak alkaline water. If pH is going down, the Zn bioavailability is increasing because its form of existence switches from Zn(OH)₄²⁻ to the good-dissolved Zn²⁺ [32].

Studies conducted in diverse regions like the upper Gulf of Thailand and English estuaries have revealed that zinc’s organic complexing capacities in interstitial waters range from 0.02 to 2 × 10⁻⁶ M, and 0.006 to 0.04 × 10⁻⁶ M in surface waters. Stability constants (log K) between 7.4 and 9.3 decrease with higher salinity. Pollution can saturate the organic complexing capacity, lowering zinc’s organic fraction, which typically comprises 93–98% in interstitial waters and 13–26% in surface waters [83].

The distribution of Zn in sediments across upstream, midstream, and downstream areas during both dry and wet seasons revealed a small amount in the residual fraction, which might be unavailable for accumulation for living organisms [12]. Conversely, the majority of Zn was found in the exchangeable, reducible, and oxidizable fractions, indicating its potential mobility and bioavailability to organisms within the Laojie River ecosystem, posing ecological risks. Notably, during the dry season, midstream and downstream locations exhibited a significant proportion of Zn in the exchangeable fraction, ranging from 24% to 47%. This suggests a predominantly anthropogenic origin for Zn, raising concerns regarding potential human exposure in these areas [12].

Sediments typically act as a source of zinc in the freshwater zone, releasing organically complexed Zn into the water column at rates of 0.3 to 15.5 μg·m⁻²·d⁻¹ through diffusion [84]. The mobility of zinc in water is largely influenced by pH, as well as factors like clay content, phosphorus availability, organic matter content, and redox conditions [85,86]. In groundwater, heavy metals in freshwater sediment generally remain immobile and do not transfer to the overlying aquatic ecosystem [87]. At low concentrations of copper and zinc, zinc tends to be retained in the solid phase, making it more difficult to extract [88].

Recent theories on the adsorption of trace metals at oxyhydroxide surfaces indicate a pH-dependent KA. However, field measurements of binding intensity often differ from equilibrium constants observed in laboratory experiments. Observations suggest that Fe-bound zinc can be mobile, with its distribution between the water column and sedimentary iron oxyhydroxides affected by pH. In shallow lakes, most mobile Zn is predicted to be in the water column at pH < 5 and in sediments at pH > 7. The pH interval, 5.5–6.5, common in poorly buffered lakes receiving acidic deposition, is where changes in water column pH significantly impact Zn partitioning [89].

Field experiments conducted in stormwater-constructed wetlands suggest a higher substrate affinity for copper over zinc [90]. Similar results have been reported in studies using sediments from shallow lakes, where the sorption capacity was found to be lowest for zinc, decreasing in the order Pb²⁺ > Cd²⁺ > Cu²⁺ > Zn²⁺ [91]. The decrease in zinc activity is attributed to the formation of metallo-organic complexes and an increase in the solution’s ionic strength. Sorption complexes within the examined profiles are primarily composed of organic matter and amorphous forms of Fe, Al, and Mn. Calcium and magnesium ions dominate the sorption complex, with sodium and potassium ions playing subordinate roles [88]. Similar findings were observed using Tessier–BCR sequential extractions. Zinc has been revealed to be predominantly associated with residual fractions and Fe/Mn oxihydroxyde fractions, while copper may be linked to carbonate minerals [90].

Non-linear fitting with FITEQL 4.0 reveals consistent conditional zinc binding constants across various sources of natural organic matter at the same pH, suggesting independence from geographical origin. Unified binding parameters adequately represent zinc–natural-organic-matter complexation. Total ligand concentrations exhibit no discernible variation with pH. However, conditional zinc-binding constants demonstrate a linear rise with pH, indicating enhanced stability of zinc–natural-organic-matter complexes at higher pH levels [92]. The adsorption of Zn²⁺ and Cd²⁺ is effectively represented by the Freundlich isotherm model, demonstrating high efficiency. In contrast, the adsorption behavior of copper differs, as the Langmuir isotherm model was deemed more appropriate [91]. Zinc behaves differently from copper and lead, undergoing ion exchange or charge compensation at the surface of sand. The sorption capacity of sand for zinc is limited to 0.5 μmol∙g⁻¹ [90].

Exploration of the remobilization mechanism of zinc in sediments remains challenging due to dynamic hydrochemical conditions. Seasonal studies using techniques such as chemical sequential extraction procedure and diffusive gradient in thin films have shown that in summer, reductive dissolution of Fe/Mn oxides drives Zn remobilization, while in winter, cation exchange reactions facilitate this process in brackish water zones with the rate of 18.9–70.7 μg·m⁻²·d⁻¹. These conditions indicate the sediment solid phase’s capacity to resupply Zn to the porewater in both seasons [84].

2.3. Zn Bioaccumulation in Biota and Toxicity

Zinc accumulation in fish tissue is believed to be an important indicator of environmental health [93] and can have far-reaching implications for aquatic ecosystems. Examining the accumulation patterns of zinc in fish tissues provides valuable insights into its potential for bioaccumulation and then bioconcentration in a trophic chain, subcellular and tissue distribution, potential toxicity, cellular and molecular targets, and ecological impact across different environmental settings.

Studies on zinc accumulation in fish tissues have revealed various patterns influenced by factors such as exposure concentration, duration, temperature, co-exposure with other metals, notably calcium, and the animal species. It is believed that metals, including zinc, may accumulate in tissues when their levels exceed the background concentration in the surrounding environment. However, these assumptions have mainly been derived from findings discovered in acute conditions. For instance, a linear uptake was evident in rainbow trout and Nile tilapia at higher zinc levels (>500 µM and 3.5–7.0 mg∙L⁻¹ correspondingly) [94]. The clear patterns disclosed in the liver and gill tissues of Nile tilapia, showing a significant increase in tissue Zn content over time. Specifically, levels rose from 25.4 ± 0.47 µg∙g⁻¹ to 109.5 ± 4.36 µg∙g⁻¹ after 6 weeks of exposure [95]. Channa punctatus exposed to acute zinc concentrations of 6.62 and 13.24 mg∙L⁻¹ for 45 days demonstrated time- and concentration-dependent accumulation in liver, kidney, and intestine tissues. Specifically, in the liver, zinc levels increased from 10.11 ± 1.08 to 67.31 ± 1.29 µg Zn∙g⁻¹ after 45 days at the higher concentration of 13.24 mg∙L⁻¹ [96].

Similar findings were observed in natural conditions in an extremely polluted environment. Gambusia holbrooki introduced into the Upper Enoree River in South Carolina, USA, exhibited a two-fold increase in zinc concentration in their gills (150 ppm∙g⁻¹ tissue compared to 77.4 ppm∙g⁻¹ tissue in fish from the reference site) and gonads (45 ppm∙g⁻¹ tissue compared to 30.7 ppm∙g⁻¹ tissue in fish from the reference site). This was observed only in the contaminated stream with the highest zinc concentration, reaching as much as 5.72 mg∙L⁻¹, alongside significant mortality. However, no significant zinc accumulation was observed in streams with lower zinc concentrations, not exceeding 1.24 mg∙L⁻¹ [97]. Similarly, common carp juveniles exposed for 10 days to metals in sublethal concentrations, including 10% and 50% of the 96 h LC₅₀, showed a positive, dose-dependent uptake for cadmium and copper, but not for zinc reaching in gills 24.39 ± 1.42 μmol∙g⁻¹ under the uptake rate of 0.05 μg∙g⁻¹ DW/h, along with a lack of correlation for zinc bioaccumulation with mortality [98]. Notably, the accumulated-zinc concentration was much higher than for both juveniles of common carp (550–680 µg∙g⁻¹ DW [99]) and moderately polluted animal habitat area (127.9–276.0 µg∙g⁻¹ FW [100].

Invertebrates, particularly mollusks, have shown a significant ability to accumulate zinc from their environment. For instance, increased water column concentrations of Zn and Cu were correlated with higher metal concentrations in the soft-body tissues of transplanted Elliptio buckleyi mussel populations, independent of their growth rate [101]. Pond snails exposed to zinc (130 µg L⁻¹) for 14 days also demonstrated clear zinc accumulation [102].

However, some studies indicate no correlation or even a reverse relationship between zinc accumulation in mollusks and environmental concentrations. For example, after 48 h, Zn levels at 6.7 μg·L⁻¹ resulted in 35% of the added Zn remaining in solution, but zinc tissue levels in Dreissena polymorpha did not significantly change. Additionally, labile zinc in mussel tissues increased, but there was no significant correlation between total and labile zinc levels [103]. Exposures of Anodonta anatina mussels from a forestry site had no effect on metal levels, whereas those from an agricultural site showed decreased Zn concentrations in the digestive gland compared to control [104]. Likewise, the intertidal gastropod Tympanotonus fuscatus L. from Nigeria showed no significant regression in Zn accumulation despite a clear correlation for Cd [105]. There was also no significant relationship between metal content in water, sediment, and the gastropod L. carinatus in Lake Manzala [106].

While zinc pollution is generally considered moderate, it is unusual for fish to accumulate significant amounts of zinc. Specifically, while zinc concentrations in the water column of the Lot River vary greatly from one station to another—Riou-Mort/Bouillac (BO) station = 17, Bouillac station/Cajarc (CA) station = 3 in the unfiltered samples—zinc bioaccumulation in the organs differs only slightly. In fact, no significant difference between the control fish and those from BO and CA can be observed in the gills, liver, muscle, and kidneys. Only in the gills of chub (Leusciscus cephalus), roach (Rutilus rutilus), perch (Perca fluviatilis), and the bream (Abramis brama) from Riou-Mort were zinc levels higher (100–300 µg g⁻¹ ww) than in the reference site (10–90 µg g⁻¹ ww) [107]. No correlation between zinc concentration in water and tissue was revealed. Also, in Barbus marequensis from the lower Olifants River, no clear trend between localities, although significant differences in zinc in different parts of the river, but significant seasonal differences were detected [108]. Similar findings were disclosed also for higher vertebrates: zinc content in frog from rural and urban ponds in Western Ukraine was showed to varied from 70.4 ± 13.6 to 221.6 ± 12.1 µg Zn g⁻¹ DW with seasonal changes were more significant than spatial [6].

The lack of clear correlation between zinc concentrations in the environment and in the tissues of animals living in zinc-enriched water bodies leads to the conclusion that zinc tissue concentration and BAF/BCF are not reliable indicators of zinc pollution. This conclusion is supported by the existing literature. For instance, McGeer et al. (2003) pointed out the limitations of using BCF as the sole indicator of zinc bioaccumulation, noting an inverse relationship between BCF and exposure concentration [109]. Cyprinus carpio habitat ponds in rural and industrial areas in Western Ukraine with low levels of industrial activity showed zinc concentrations ranging from 86.3 ± 22.1 µg∙g⁻¹ to 603.3 ± 91.7 µg Zn∙g⁻¹ FW. No correlation between water and accumulated zinc in tissues was detected [110].

Given the evidence, it is highly probable that fish develop effective strategies to avoid accumulating metals, even in environments with high zinc levels. Fish likely utilize their olfactory systems and mucus-like secretions to avoid metal accumulation [111,112]. Additionally, they employ efficient homeostatic mechanisms, such as the use of metallothioneins, to regulate zinc concentrations within their bodies. Metallothioneins are ubiquitous, metal-binding proteins characterized by their sulfur-rich and highly dynamic nature, serving various roles such as metal buffering, detoxification, and sequestration of reactive oxygen species [82]. It is probable that metallothioneins primarily function in the regulation of zinc metabolism [113], thereby preventing significant zinc accumulation under stressful conditions, except in cases of extreme pollution. For instance, in carp from zinc-enriched ponds in industrial sites, lower levels of zinc were observed in the liver alongside higher levels of zinc bound to metallothioneins. Similarly, Zn-exposed mussels from reference ponds in forest areas showed a three-fold elevation in the metal-binding capacity of metallothioneins while maintaining stable tissue zinc levels in the digestive gland of Anodonta anatine [104]. An increase in metallothionein-bound zinc was also noted in zinc-exposed Oreochromis niloticu [114]. Similar results, highlighting the involvement of metallothioneins in coping with zinc pollution, have been observed in studies involving transplanted freshwater species [115]. These observations emphasize the importance of metallothioneins in mitigating the effects of zinc pollution in freshwater species.

Another potential mechanism contributing to the ability of fish to withstand zinc pollution and prevent overaccumulation involves zinc-induced mucus secretion in their intestines. This mechanism acts as a regulator of zinc uptake, facilitating absorption at lower concentrations while limiting it at higher levels. In experiments involving rainbow trout exposed to either seawater or 20 ppm zinc, researchers observed heightened mucus secretion, evidenced by comparable sialic acid content regardless of the environment. However, the production of sialic acid, and consequently mucus, was notably elevated under seawater and zinc-exposed conditions compared to freshwater environments [116].

Moreover, zinc homeostasis can be maintained through lysosomal regulation even at low exposure levels. Research conducted on fish fin cell cultures exposed to zinc up to 250 μM demonstrated a significant increase in lysosomal number and size, accompanied by enhanced lysozyme activity, indicating detoxification responses involving zinc influx and storage. Additionally, this study emphasizes the interaction between lysosomes, responsible for zinc accumulation, and mitochondria, which are the primary targets of free zinc in cases of spillover [117].

Although no significant Zn accumulation in animal tissues is expected, Zn is able to cause significant molecular and cellular adverse outcomes in a wide range of water habitants. It has been shown in a wide range of organisms that a heightened level of Zn in the environmental milieu is followed by the appearance of oxidative stress determined as higher levels of lipid peroxidation and reactive oxygen species overproduction, mitochondrial dysfunction, immune disorders, apoptosis, and depletion of energy reserves [7,104,117,118].

To sum up, the relationship between environmental zinc concentrations and zinc accumulation in water animals is not always straightforward, with some research indicating no correlation or even an inverse relationship. Various adaptive mechanisms, such as metallothionein expression, scale-mediated chelation, mucus secretion, and lysosomal storage, are employed by these organisms to counteract zinc exposure, thereby mitigating its toxic effects in aquatic ecosystems. Consequently, zinc accumulation in freshwater animals may not serve as a reliable biomarker of zinc water pollution due to the potential for metabolic regulation.

2.4. Bioremediation of Zinc Pollution

Bioremediation of zinc pollution in freshwater ecosystems has caught significant attention due to its cost-effectiveness and environmental benefits. As one form of bioremediation, phytoremediation involves using plants to remove contaminants from water and soil. One study explored the use of marigold plants for zinc removal and reported effective accumulation of zinc in its roots, stems, and leaves, demonstrating its potential for treating zinc-contaminated sites [119]. Also, water spinach was tested for phytoremediation of Zn-, Cu-, and Pb-contaminated Tempe lake, showing a promising relationship between the phytoremediation time and the BCF value. Infra-Red (IR) data proved metal binding in plants with the functional groups C=S, C=N, and OH [120]. Another plant, Callitriche cophocarpa, demonstrated significant capabilities to accumulate metals in its shoots from water in the following order: Zn (1120) < Tl (251) < Cd (71) < Pb (35) (mg·kg⁻¹ DW) [121]. Later findings however expanded on this, providing that roots may have greater accumulation potential for most metals namely Cd, Co, Cr, Cu, Fe, Mn, Ni, and Zn than shoots, indicating restricted mobility and translocation by C. cophocarpa to shoots [122]. To increase the surface of contact and facilitate direct contact between plant roots and contaminated water, floating treatment wetlands are being used, thereby enhancing zinc adsorption and deposition [123].

The presence of zinc in water alongside other inorganic elements or organic substances plays a significant role in its absorption by plants and microorganisms [124]. Elements like silicon can modify zinc aggregation within plant cells, enhancing its uptake [125]. In contrast, chloride ions can bind with zinc, reducing its removal efficiency. Additionally, organic contaminants can have an antagonistic effect, inhibiting the extraction of metals from the soil by plants [126]. On the other side, the complexation that occurs when these substances coexist can decrease the toxicity of dissolved metal ions [127], including zinc, which is beneficial for managing metal toxicity in aquatic animals. Among promising materials appears chitosan [128]. It not only reduces metal toxicity but also serves as an effective adsorbent for remediation, capable of removing over 70% of metallic ions [129].

So, it has to be mentioned that adsorption is a widely explored method for mitigating metal pollution in water besides phytoremediation. Instead of conventional adsorbents like activated charcoal, different natural materials, including seaweed, microbial cell walls, exopolysaccharides, extracellular proteins, mushroom residue, pine bark, chitosan, and leaves seem to be promising adsorbents [130]. In this context, the biosorption of Zn from aqueous solutions by living Pseudomonas aeruginosa was investigated. Factors like pH, bacterial dosage, initial Zn(II) concentration, contact time, and temperature significantly influenced the removal process. The maximum uptake capacity of P. aeruginosa for Zn(II) ions was 46.1 mg/g under these conditions. The second-order kinetic model and Freundlich isotherm model best described the biosorption data [131]. Further studies investigated the biosorption of Pb(II) and Zn(II) using a metal-tolerant bacterium, Oceanobacillus profundus KBZ 3-2, isolated from a contaminated site. The results showed that the maximum removal percentage for Zn(II) was 54% at an initial concentration of 2 mg∙L⁻¹, achieved at pH 6 and 30 °C, whereas the maximum removal percentage for Pb(II) was 97% at an initial concentration of 50 mg∙L⁻¹. The isolated bacteria sequestered metals in the extracellular polymeric substance, which facilitates ion exchange and metal chelation–complexation through ionizable functional groups such as carboxyl, sulfate, and phosphate present in proteins and polysaccharides [132]. Additionally to microorganisms, leaves of Inula viscosa were tested as an adsorbent, and the Zn biosorption process was described by the Langmuir isotherm and pseudo-second-order kinetic models, revealing spontaneous and endothermic Zn²⁺ sorption on leaves [133].

3. Copper

3.1. Occurrence of Copper in Surface Waters and Sediments

In aquatic systems, copper frequently originates from both natural and anthropogenic sources. Natural sources encompass geological deposits, volcanic activities, and the weathering and erosion of rocks and soils. On the anthropogenic side, activities such as mining, metal and electrical manufacturing, agriculture through the application of pesticides and fertilizers, and effluents from wastewater treatment plants significantly contribute to the elevated copper levels found in water bodies [134]. These activities have profound impacts on copper concentrations in rivers and lakes, causing notable spatio-temporal trends. For instance, rivers downstream of major urban industrial areas often show increased metal concentrations. Additionally, former mining areas situated in crystalline zones continue to release heavily contaminated sediments over decades. Conversely, since the 1970s, there has been a general decline in metal concentrations, although some metals, such as chromium and nickel, have remained stable at lower levels [135].

Copper is present in both the water column and sediments of inland waters, with average concentrations varying significantly across Europe, Asia, and America, reflecting diverse sources and environmental conditions. Notably, surface water generally contains lower copper concentrations compared to sediments. In European water bodies, copper concentrations in surface water typically range from 0.2 to 5.5 µg∙L⁻¹ (Table 1 and Table 2), primarily due to strict regulations and effective pollution control measures. For example, in the Mała Wełna River system (Western Poland), surface-water copper concentrations averaged 0.004 ± 0.015 mg∙L⁻¹, with a peak of 0.089 mg∙L⁻¹ [136]. This is comparable to the Oder River in Poland (0.0235 mg∙L⁻¹) and the Warta River near Mosina (0.007 mg∙L⁻¹) [137,138]. Similar levels are noted in Ukrainian water bodies, specifically in the river of the Dnister basin in West Ukraine, where copper concentration reached 39.4 µg∙L⁻¹ [110]. Occasionally, higher copper levels are observed, such as in the Krzna River (Poland), ranging from 0.5 to 1.4 mg∙L⁻¹ [29], highlighting the influence of agrotechnical activities. In Mirgenbach Lake (Northeastern France), surface-water copper concentrations averaged 38 µg∙L⁻¹ (ranging between 24 and 62 µg∙L⁻¹), with sediment concentrations between 656 and 1223 mg∙kg⁻¹, averaging at 912 mg∙kg⁻¹ [139].

In Asia, surface-water copper concentrations vary widely, reflecting disparate environmental conditions and levels of industrialization. For example, on the southeastern coast of Andhra Pradesh, India, concentrations can be as low as <0.027 µg∙L⁻¹ [36], whereas in Thailand’s Tha Chin River, levels can reach as high as 1200 µg∙L⁻¹ [39]. This stark contrast underscores the significant anthropogenic impacts from industrial and agricultural activities. In the Chu-Talas Basin in Central Asia, copper concentrations range from 0.154 to 11.853 µg∙L⁻¹, with an average of 2.210 µg∙L⁻¹ [78]. Similarly, in the Laojie River, copper concentrations range from 1 to 43 µg∙L⁻¹, along with other metals such as arsenic (2–4.3 µg∙L⁻¹), zinc (9–224 µg∙L⁻¹), and nickel (0–2.7 µg∙L⁻¹) [12].

In North America, surface-water copper concentrations are generally moderate. In the Alaska riverine system, copper concentrations vary from 2 to 260 µg∙L⁻¹ [140]. In the upper Columbia River, concentrations typically fall below U.S. water quality criteria designed to protect aquatic life, except at the sediment–water interface, where levels can reach 24 µg∙L⁻¹, exceeding the criteria [141]. In British Columbia, Canada, the median concentration of total copper across multiple stations is <1 µg∙L⁻¹, with a median of 0.68 µg∙L⁻¹ and a 90th percentile value of 3.0 µg∙L⁻¹ [142], indicating relatively low contamination levels.

In Africa, recorded copper concentrations in surface waters, such as Lake Naivasha (Kenya) and the Nzhelele River (South Africa), range between 0.0257 mg∙L⁻¹ and 0.066 mg∙L⁻¹, suggesting potential influences from agriculture, particularly horticulture and floriculture, as well as contributions from rivers traversing agricultural areas (Mutia and Edokpayi). Despite these sources, copper levels remain below Kenyan safety standards for water usage, which stand at 2 mg∙L⁻¹ for general use [142] and 1 mg∙L⁻¹ as the maximum allowable level for effluent discharge into the environment [143]. However, copper levels in Nigeria’s Abonnema shoreline range from <0.001 to 0.0772 mg∙L⁻¹, slightly higher than in Kenya [144], likely due to industrial discharge, urban runoff, and different agricultural practices.

In terms of sediment concentrations, Europe’s sediments typically exhibit copper levels ranging from 50 to 150 mg∙kg⁻¹. However, industrial areas can have significantly higher concentrations. For instance, sediments in France’s Lot River have copper levels between 40,600 and 264,000 µg∙kg⁻¹ (Table 3), primarily due to industrial activities and agricultural runoff. The River Miño in the Iberian Peninsula, affected by vineyards, shows sediment copper concentrations from 18 to 209 mg/kg [145]. The OSPAR assessment shows sediment copper levels varying widely, from 1 mg∙kg⁻¹ to 355 mg∙kg⁻¹, with a median value of 22 mg∙kg⁻¹ [146].

Asian sediments show significantly high copper concentrations often between 100 and 300 mg∙kg⁻¹ due to extensive mining and rapid urbanization. Modified degree of contamination (mCd) calculated from contamination factor in certain regions exceed mCd > 1.5 [147], indicating substantial anthropogenic impacts. In Chinese sediments, levels vary enormously, ranging from 8 to 753 mg∙kg⁻¹, with some areas showing up to 146,200 µg∙kg⁻¹ [44]. This high variability and elevated maximum concentrations underscore the significant impact of riverine inputs and industrial activities in these regions. Although no comprehensive assessment of time trends was conducted, it is evident that continuous industrial discharge and urban runoff are primary contributors to the alarmingly high levels of metals in sediments, water, and biota.

This indicates substantial environmental pressure from anthropogenic activities. The sediment samples collected over a five-year period showed consistent heavy metal concentration fingerprints, suggesting that natural attenuation processes were insufficient. Consequently, intervention through advanced treatment technologies is necessary to mitigate these high contamination levels and improve sediment quality.

In North America, sediment copper concentrations typically range between 30 and 100 mg∙kg⁻¹, reflecting varying degrees of industrial activity and urban development, as areas closer to industrial sites and urban centers tend to have higher concentrations. The 2014 Mount Polley copper–gold mine tailings spill in British Columbia resulted in maximum copper concentrations in the affected sediments around 410 mg∙kg⁻¹, far exceeding the Canadian sediment quality guidelines for the protection of aquatic organisms (197 mg∙kg⁻¹) [148].

African sediments also reflect varying copper concentrations, demonstrating diverse levels of metal accumulation. In Kenya’s Malewa and Karati Rivers, sediment concentrations range from 3.00 to 8.48 mg∙kg⁻¹ [149]. The generally low concentrations of Cu in these sediment samples are attributed to their dissolved or suspended states in the water, facilitated by water turbulence, or their uptake by aquatic plants. In comparison, sediment samples from the Nzhelele River in South Africa reflected Cu concentrations ranging from 2.182 to 566 mg∙kg⁻¹, with in general moderate ecological risk levels (Er = 62.90). However, a significant spike to 566 mg∙kg⁻¹ was implied potential episodic contamination events [150]. Further south in the Molopo River, Mahikeng, South Africa, sediment samples revealed Cu concentrations varying from 15.07 to 170.9 mg∙kg⁻¹, with Zn also flagged as a significant contaminant (16.52–349.8 mg∙kg⁻¹). Contamination factors (CF) for Cu ranged from moderate to considerable, indicating varying degrees of contamination. The enrichment factor (EF) for Cu suggested both moderate (2 EF < 5) and substantial (5 EF < 20) enrichment levels, indicative of substantial anthropogenic influence (EF > 2) [151].

To sum up, copper contamination in surface waters and sediments is widespread, originating from both natural and anthropogenic sources. The variability in concentrations across different regions highlights the influence of local environmental conditions and human activities. European countries benefit from strict regulations resulting in lower contamination levels, while industrial and agricultural activities in Asia and Africa contribute to higher copper concentrations. In the Americas, contamination levels are moderate but can spike due to specific events like mining spills. Sediments typically exhibit higher copper concentrations than the water column, with particularly high levels near industrial sites. Ongoing monitoring and mitigation efforts are essential to manage and reduce copper contamination, saving aquatic ecosystems and ensuring compliance with regulatory standards.

Table 3. Occurrence of copper in surface waters and sediments.

	Country	Place	Concentration Range: Surface Water (µg·L−1) Sediment (µg·kg−1 D.M.)	Ref.
Cu	ASIA
	India	Southeastern coastal: Andhra Pradesh	<0.027	Mahdi et al., 2022 [36]
	Thailand	Chao Phraya River (n = 16)	1.81–5.63 (rainy season); <LOQ–5.62 (dry season)	Niampradit et al.2024 [37]
		Chao Phraya River (n = 9)	1.12–14.22	Chanpiwat and Sthiannopkao 2013 [38]
		Tha Chin River (n = 38)	10–1200	Veschasit et al., 2012 [39]
	Cambodia	Tonle Sap-Bassac River (n = 11)	0.25–1.62	Veschasit et al., 2012 [39]
	Indonesia	Citarum River (n = 10)	0.51–6.94	Veschasit et al., 2012 [39]
		Winongo River (n = 8)	0–40	Fadlillah et al., 2023 [40]
	Malaysia	Linggi River (n = 15)	0.06–3.06	Razak et al., 2021 [41]
		Semenyih River (n = 8)	0.84–7.33	Al-Badaii 2014 [42]
	Vietnam	Saigon River (n = 8)	0.55–16.51	Chanpiwat and Sthiannopkao 2013 [38]
	Iran	Chah Nimeh reservoir—surface water (n = 7)	11–115 (spring) 174–217 (summer)	Bazrafshan et al., 2015 [43]
		Chah Nimeh reservoir—sediment (n = 7)	3000–5000 (spring) 3400–5180 (summer)
	China	Dianchi Lake: water surface/sediment	1.36/146,200	Liu et al., 2021 [44]
		Chaohu Lake: water surface/sediment	2.56/26,900	Wang et al., 2016 [45]
		Daye Lake: water surface/sediment	2.16–3.01/143,100	Wang et al., 2023 [46]
		Yangtze River: water surface/sediment	2.86/28,500	Li et al., 2020 [47]
	EUROPE
	Poland	Muchawka River (n = 16/12): surface water/sediment	0.2–1.8/700–4300	Kluska and Jabłonska 2023 [48]
		Liwiec River (n = 32) surface water/sediment	0.3–1.6/900–7400
	Greece	Bay and Gulf of Thessaloniki (Aegean Sea including the Bay and Gulf of Thessaloniki)	0.8–5.5	Christophoridis et al., 2009 [49]
	Turkey	Ataturk lake—sediment	14,570	Karadede and Unlu, 2000 [50]
		The Dipsiz stream (sediment)—tributary of the river Buyuk Menderes	13.000 ± 9.000	Demirak et al., 2006 [51]
		The Dipsiz stream—tributary of the river Buyuk Menderes	0.365 ± 0.394
	Spain	Guadaira river: surface water/sediment	10–20/13,900–142,600	Enguix Gonzalez et al., 2000 [52]
		Tinto River (sediment)	180,000–2650.000	Galan et al., 2003 [52]
	France	Cajarc site, Lot River (sediment)	40,600–264,000	Audry et al., 2004 [54]
	Germany	Malter Reservoir	<240,000	Muller et al., 2000 [55]
	Scotland	Lochnagar (sediment)	8000–25,000	Yang et al., 2002 [56]
	Switzerland	Lake Zurich (sediment)	20,000–78,000	Von Gunten et al., 1997 [57]
	Netherlands	Meuse River (sediment)	50,000–105,000	Van der Berg et al., 1999 [58]
	AMERICA
	Mexico	Campeche Bay	5.6–11.4	Villanueva 2000 [152]
		Continental shelf—Campeche (sediment)	3800–18,700	Macias et al., 1999 [153]
		Continental shelf—Tamaulipas (sediment)	3200–24,980	Ponce 1995 [59]
		Continental shelf— Veracruz (sediment)	1600–91,250
	United States	300 coastal and estuarine sites (sediment)	35,000 (Average)	O’Connor and Cantillo 1992 [61]
		San Francisco Bay	0.2–5.3	Sadiq 1992 [154]
	U.S. Virgin Islands	Developed sites/undeveloped sites (wetland, forest, shrub, rangeland)	27,100–82,900/31,700–93,000	Lancellotti et al., 2023 [62]
	Urugway	Montevideo Harbor (sediment)	59,000–126,000 (summer) 59,000–135,000 (winter)	Muniz et al., 2004 [63]
	Brazil	Jurujuba SOund	5000–213,000	Baptista et al., 2000 [64]
	Venezuela	Coral reef sediment	6000–40,000	Bastidas et al., 1999 [65]
	Chile	Southern Fjords	16,000–22,000	Ahumada et al., 2015 [66]
	AFRICA
	Uganda	Kampala and Mbarara districts	0.034–0.146 (dry season) 0.235–0.322 (wet season)	Sanusi et al., 2024 [155]
	Malawi	Lake Chilwa: surface water/sediment	BDL–47.83/20,020	Mussa et al., 2019 [67]
	Cameroon	Municipal Lake (sediment)	42,800–142,000	Ekengele et al., 2008 [68]
	Egipt	El-Mex Bay	3690–4900	Abdallah 2008 [69]
	Morocco	Sebou Estuary sediments	51,500	Cheggour et al., 2005 [70]
		Nador lagoon	4000–1,190,000	Bloundi et al., 2008 [71]
	Zambia	Kafue River (sediment)	12,855	Pettersson and Ingri 2001 [156]
	Namibia	sediments of the Gruben River	10,500.000	Taylor and Kesterton 2002 [73]
	Algeria	Tafna Wadi	10–50	Benmostefa et al., 2022 [74]
	Tanzania	Mwanza Gulf of Lake Victoria (sediment)	26,100	Kishe and Machiwa 2003 [75]

Table 3. Occurrence of copper in surface waters and sediments.

	Country	Place	Concentration Range: Surface Water (µg·L−1) Sediment (µg·kg−1 D.M.)	Ref.
Cu	ASIA
	India	Southeastern coastal: Andhra Pradesh	<0.027	Mahdi et al., 2022 [36]
	Thailand	Chao Phraya River (n = 16)	1.81–5.63 (rainy season); <LOQ–5.62 (dry season)	Niampradit et al.2024 [37]
		Chao Phraya River (n = 9)	1.12–14.22	Chanpiwat and Sthiannopkao 2013 [38]
		Tha Chin River (n = 38)	10–1200	Veschasit et al., 2012 [39]
	Cambodia	Tonle Sap-Bassac River (n = 11)	0.25–1.62	Veschasit et al., 2012 [39]
	Indonesia	Citarum River (n = 10)	0.51–6.94	Veschasit et al., 2012 [39]
		Winongo River (n = 8)	0–40	Fadlillah et al., 2023 [40]
	Malaysia	Linggi River (n = 15)	0.06–3.06	Razak et al., 2021 [41]
		Semenyih River (n = 8)	0.84–7.33	Al-Badaii 2014 [42]
	Vietnam	Saigon River (n = 8)	0.55–16.51	Chanpiwat and Sthiannopkao 2013 [38]
	Iran	Chah Nimeh reservoir—surface water (n = 7)	11–115 (spring) 174–217 (summer)	Bazrafshan et al., 2015 [43]
		Chah Nimeh reservoir—sediment (n = 7)	3000–5000 (spring) 3400–5180 (summer)
	China	Dianchi Lake: water surface/sediment	1.36/146,200	Liu et al., 2021 [44]
		Chaohu Lake: water surface/sediment	2.56/26,900	Wang et al., 2016 [45]
		Daye Lake: water surface/sediment	2.16–3.01/143,100	Wang et al., 2023 [46]
		Yangtze River: water surface/sediment	2.86/28,500	Li et al., 2020 [47]
	EUROPE
	Poland	Muchawka River (n = 16/12): surface water/sediment	0.2–1.8/700–4300	Kluska and Jabłonska 2023 [48]
		Liwiec River (n = 32) surface water/sediment	0.3–1.6/900–7400
	Greece	Bay and Gulf of Thessaloniki (Aegean Sea including the Bay and Gulf of Thessaloniki)	0.8–5.5	Christophoridis et al., 2009 [49]
	Turkey	Ataturk lake—sediment	14,570	Karadede and Unlu, 2000 [50]
		The Dipsiz stream (sediment)—tributary of the river Buyuk Menderes	13.000 ± 9.000	Demirak et al., 2006 [51]
		The Dipsiz stream—tributary of the river Buyuk Menderes	0.365 ± 0.394
	Spain	Guadaira river: surface water/sediment	10–20/13,900–142,600	Enguix Gonzalez et al., 2000 [52]
		Tinto River (sediment)	180,000–2650.000	Galan et al., 2003 [52]
	France	Cajarc site, Lot River (sediment)	40,600–264,000	Audry et al., 2004 [54]
	Germany	Malter Reservoir	<240,000	Muller et al., 2000 [55]
	Scotland	Lochnagar (sediment)	8000–25,000	Yang et al., 2002 [56]
	Switzerland	Lake Zurich (sediment)	20,000–78,000	Von Gunten et al., 1997 [57]
	Netherlands	Meuse River (sediment)	50,000–105,000	Van der Berg et al., 1999 [58]
	AMERICA
	Mexico	Campeche Bay	5.6–11.4	Villanueva 2000 [152]
		Continental shelf—Campeche (sediment)	3800–18,700	Macias et al., 1999 [153]
		Continental shelf—Tamaulipas (sediment)	3200–24,980	Ponce 1995 [59]
		Continental shelf— Veracruz (sediment)	1600–91,250
	United States	300 coastal and estuarine sites (sediment)	35,000 (Average)	O’Connor and Cantillo 1992 [61]
		San Francisco Bay	0.2–5.3	Sadiq 1992 [154]
	U.S. Virgin Islands	Developed sites/undeveloped sites (wetland, forest, shrub, rangeland)	27,100–82,900/31,700–93,000	Lancellotti et al., 2023 [62]
	Urugway	Montevideo Harbor (sediment)	59,000–126,000 (summer) 59,000–135,000 (winter)	Muniz et al., 2004 [63]
	Brazil	Jurujuba SOund	5000–213,000	Baptista et al., 2000 [64]
	Venezuela	Coral reef sediment	6000–40,000	Bastidas et al., 1999 [65]
	Chile	Southern Fjords	16,000–22,000	Ahumada et al., 2015 [66]
	AFRICA
	Uganda	Kampala and Mbarara districts	0.034–0.146 (dry season) 0.235–0.322 (wet season)	Sanusi et al., 2024 [155]
	Malawi	Lake Chilwa: surface water/sediment	BDL–47.83/20,020	Mussa et al., 2019 [67]
	Cameroon	Municipal Lake (sediment)	42,800–142,000	Ekengele et al., 2008 [68]
	Egipt	El-Mex Bay	3690–4900	Abdallah 2008 [69]
	Morocco	Sebou Estuary sediments	51,500	Cheggour et al., 2005 [70]
		Nador lagoon	4000–1,190,000	Bloundi et al., 2008 [71]
	Zambia	Kafue River (sediment)	12,855	Pettersson and Ingri 2001 [156]
	Namibia	sediments of the Gruben River	10,500.000	Taylor and Kesterton 2002 [73]
	Algeria	Tafna Wadi	10–50	Benmostefa et al., 2022 [74]
	Tanzania	Mwanza Gulf of Lake Victoria (sediment)	26,100	Kishe and Machiwa 2003 [75]

3.2. Copper Mobility and Complexation

The copper absorption–desorption in freshwater reservoirs is a complex process influenced by the interaction of sediment composition, water chemistry, and environmental factors. It is believed that the mobility of metals is primarily determined by the mineralogical composition of sediments, which impacts parameters such as specific surface area and cation exchange capacity, both of which are crucial in the adsorption of metals [157].

Natural water sediments include minerals, with clay minerals such as illite, kaolinite, and montmorillonite forming the core components. These clay particles frequently contain metal oxides such as iron oxide, alumina, and manganese oxide, as well as organic matter primarily consisting of humic acid and tannic acid. These elements have a vital role in the adsorption process of metal contaminants such as copper [158]. Quartz, a prevalent non-clay mineral, has been found to have a significant adverse impact on metal exchange. It enhances the release of metals from sediments into the water in riverine systems. In contrast, clays and mineral oxides are more effective in estuarine environments, as they enhance the movement of metals from the water column to the sediments [157].

Particulate organic matter has significant effects on the adsorption of metals in river environments. In contrast, in estuarine ecosystems, particulate phosphorus assumes a crucial role. The movement of copper is frequently controlled by alkaline conditions and moderate levels of dissolved organic carbon, which promote the formation of complexes and decrease the activity of free copper ions. Elevated levels of inorganic anions, such as chloride and sulfate, can result in the creation of soluble copper complexes, which in turn increases the mobility of copper within the water column.

The adsorptive capacity of sediments can vary significantly, impacting the movement of copper. The sorption capacity of sediments for metal ions varies, with the highest capacity observed for lead (Pb²⁺), followed by cadmium (Cd²⁺), copper (Cu²⁺), and zinc (Zn²⁺). Copper ions (Cu²⁺) are usually absorbed according to the Langmuir isotherm model, which indicates a consistent surface with a limited number of adsorption sites. Conversely, the Freundlich isotherm model is more suitable for explaining the adsorption of metals such as zinc (Zn²⁺), lead (Pb²⁺), and cadmium (Cd²⁺). This suggests that the surface involved in the adsorption process is heterogeneous, with different affinities for adsorption [91]. At high initial concentrations, copper shows modest quantities of desorption, indicating unique kinetic behaviors [159].

The spatial variability in adsorption capacity underscores the influence of sediment composition on copper mobility. An analysis of Raisin River sediments reveals linear dependencies on soluble metal concentration, with adsorption densities ranging from 6000 to 9000 μg∙g⁻¹ within 48 h and partition coefficients for copper, averaging around 50 L∙g⁻¹. Interestingly, copper’s desorption behavior is moderate compared to metals like zinc and chromium. The reversible-resistant equilibrium models commonly applied fail to fully elucidate the observed phenomena, suggesting more complex kinetic interactions at play [159].

In water reservoirs such as Dianchi Lake, the pattern of copper adsorption does not necessarily align closely with variations in the amount of sediment components. Conversely, the interactions among different sediment components can generate compound effects that impact the overall adsorption capacity. The copper adsorption by different sediment treatments exhibits a stronger correspondence with the Freundlich isotherm model, highlighting the intricate quantitative associations between organic and inorganic compounds in sediments and their combined influence on the adsorption of heavy metals [158].

Copper complexation in water bodies is a sophisticated and dynamically balanced process that significantly impacts the bioavailability and toxicity of copper. Predominantly involving natural organic matter, humic and fulvic acids act as primary ligands, efficiently binding with copper and thus controlling its environmental fate. The organically complexed forms of copper are considerably less toxic compared to the free metal ion or inorganically associated species such as Cu²⁺, Cu(CO₃)₂²⁻, and Cu(OH)⁺. This reduced toxicity is linked to the lower bioavailability of organic forms, which is an outcome of their decreased chemical reactivity in comparison to inorganic copper species [160].

The copper-complexing capacity (CC) of various surface waters ranges broadly between 40 and 500 nM [160]. It has been suggested that organic-matter–lead-copper complexation overwhelmingly predominates over free or inorganic copper forms in such environments. For instance, in Swiss lakes and rivers, the total dissolved copper to free copper ion ratios ([Cu]T/[Cu²⁺]) span from 1 × 10⁵ to greater than 1 × 10⁷ at pH [161]. Interestingly, the extent of copper complexation in river waters is generally lower than in eutrophic lakes, despite similar Dissolved Organic Carbon (DOC) levels [161]. In limnological systems across Europe, humic substances highly likely to be the most prevalent ligands, with average DOC concentrations ranging from 2 to 10 mg∙L⁻¹.

Contrary to the traditional belief that humic substances are primarily responsible for trace metal complexation, this does not hold for copper in English rivers [160]. In eutrophic lakes, stronger copper complexation occurs particularly within euphotic layers when compared to oligotrophic lakes or various rivers. These observations suggest the presence of low-level but highly specific ligands in lakes, predominantly linked to high algal productivity [161]. Such copper ligands may be either directly produced by living algae releasing extracellular products or by decaying algae-leaching intracellular materials and debris. In rivers, fulvic and humic acids likely play a more significant in binding copper compared to lakes where specific strong ligands prevail. Corresponding studies in British Columbia waters correlate ligand concentrations positively with chlorophyll levels, suggesting marine phytoplankton significantly contribute to Cu ligands [162]. This indicates both marine humic substances and terrestrial organic matter contribute notably to the pool of ligands detected by immobilized Metal Affinity Chromatography, which includes smaller L2 class ligands [162].

In freshwater systems, Cu²⁺ is the predominant copper ion form, whereas in marine environments, Cu⁺ is more prevalent. Titrations of seawater with added copper (I)-binding ligands reveal that these ligands are detected upon seawater titration with copper (II), which is reduced to copper(I) within a span of 2 to 40 min, depending on the ligand’s nature. Common marine thiols such as glutathione indicate that detected ligands in seawater, previously assumed to bind copper (II), may in reality be complexing with copper(I), thus stabilizing this oxidation state in seawater [163].

Research conducted by Paul et al. (2021) using competitive ligand equilibration–adsorptive stripping voltammetry in deep-sea pore waters of the Pacific Ocean’s polymetallic nodule provinces, namely Peru Basin and Clarion–Clipperton-Zone, it was found that dissolved Cu concentrations (3 to 96 nM) generally exceeded bottom water concentrations (4–44 nM). Processes such as deep-sea mining seem unlikely to release toxic Cu²⁺ concentrations into seawater since more than 99% of Cu was found organically complexed in pore waters, with Cu²⁺ levels remaining below 6 pM for 8 out of 9 samples [164].

Unlike zinc, where Europe has the highest average levels in water bodies, the concentration of copper declined in the following order: Asia > Africa > South America > North America > Europe (Table 3). Notably, copper concentrations never exceeded WHO and USEPA standards (Table 1).

Overall, copper contamination in surface waters and sediments is widespread, originating from both natural and anthropogenic sources. The variability in concentrations across different regions highlights the influence of local environmental conditions and human activities. European countries benefit from strict regulations, resulting in lower contamination levels, while industrial and agricultural activities in Asia and Africa contribute to higher copper concentrations. In the Americas, contamination levels are moderate but can spike due to specific events, like mining spills. Sediments typically exhibit higher copper concentrations than the water column, with particularly high levels near industrial sites.

3.3. Cu Bioaccumulation in Biota and Toxicity

Cu is element necessary for the proper functioning of aquatic animals, and their uptake is regulated homeostatically through the operation of various mechanisms depending on the chemical composition of water. High concentrations of Cu are toxic to both microorganisms, plants, and higher organisms, including humans. Currently, there has been a significant increase in exposure to Cu around the world as a result of anthropogenic activities (mining, industry, and agriculture) and improper waste management. This metal is difficult to reclaim due to its relatively easy transport and high mobility, which makes it one of the most common metallic substances that contribute to environmental pollution [165,166].

According to the assumptions, in fresh water, fish absorb water-borne metals primarily through the gills, while, in salt water, they absorb them through both the gills and the intestines. Importantly, other aquatic organisms, such as phytoplankton, which are present in both fresh and salt water, can take up significant amounts of Cu from the water column and transfer the metals to other aquatic organisms in the food chain. The ability to accumulate copper is significantly dependent on seasonal variations and the developmental stages of organisms. For instance, in fish, the accumulation of metals reaches a steady state after a certain age [167]. A study by Mert et al. (2014) highlighted significant differences in metal content in fish such as tench, pike-perch, and common carp caught in Lake Damsa (Nevşehir) across four different seasons. Findings revealed that seasonal fluctuations significantly affect metal accumulation, presumably due to increased human activity and variations in fish dimensions and weights. For example, the average weight of pike-perch varies markedly from 146.4 ± 31.341 g in autumn to 775.75 ± 528.297 g in spring. The Cu content in tench ranges from the lowest values of 0.1934 ± 0.117 mg∙kg⁻¹ in winter to the highest of 15.422 ± 6.508 mg∙kg⁻¹ in autumn. In pike-perch, it ranges from 0.1094 ± 0.063 mg∙kg⁻¹ in winter to 14.0216 ± 12.993 mg/kg in autumn, while in carp, it varies from 0.1543 ± 0.116 mg∙kg⁻¹ in autumn to 1.0408 ± 1.333 mg∙kg⁻¹ in winter [168].

Although copper is an essential nutrient for fish, necessary to maintain basic metabolic processes, elevated concentrations can lead to adverse effects in several key physiological pathways. Research has consistently shown that aquatic organisms, particularly fish and crustaceans, are 10 to 100 times more sensitive to copper toxicity than mammals [167,169,170]. In Polish freshwater environments, including natural reservoirs and farm ponds, among them lakes Roś, Nidzkie, Ukiel, Czos, Śniardwy, and Niegocin, as well as Barycz River, the Cu concentrations in fish varied from 0.3 ± 0.2 mg∙kg⁻¹ to 1.3 mg∙kg⁻¹. The lowest mean Cu content was found in pike-perch (0.03 ± 0.1 mg/kg), while the highest levels were observed in roach (0.5 ± 0.2 mg∙kg⁻¹), salmon (0.58 ± 0.41 mg∙kg⁻¹), and gilthead seabream [171,172,173,174].

Similarly, in Western Ukraine, the Cu content in the liver and gills of Cyprinus carpio from different sites varied, with levels ranging from 4.3 ± 0.3 to 9.4 ± 1.2 µg∙g⁻¹ and 1.6 ± 0.3 to 3.0 ± 0.2 µg∙g⁻¹, respectively. Seasonal pollution patterns suggest lower Cu concentrations in summer due to its deposition during spring and autumn [110].

In the large, eutrophic Lake Taihu (Meiliang Bay) in China, studies have documented fluctuating copper concentrations in fish tissues over time. Specifically, the mean Cu concentration in the skin and gonads of Cyprinus carpio ranged from 0.480 ± 0.090 mg∙kg⁻¹ to 1.152 ± 0.101 mg∙kg⁻¹, and from 4.615 ± 0.009 mg/kg to 13.617 ± 0.128 mg∙kg⁻¹ in C. auratus [175]. These levels are notably higher than those observed in fish from European freshwater habitats [110,171,172,173,174].

Recent data indicate a positive trend in reducing copper pollution in Lake Taihu over the past decade, particularly with significantly lower copper concentrations now being recorded in fish tissues from Meiliang Bay. For example, the mean concentration of Cu in the tissue samples of Cyprinus carpio and Pelteobagrus fluvidraco from Meiliang Bay, Lake Taihu ranged between 0.037 and 0.316 mg∙kg⁻¹ during the summer and 0.017–0.144 mg∙kg⁻¹ in the winter [176]. This improvement is likely attributed to China’s enhanced regulatory efforts, which now align more closely with the standards established by the European Union (EU) and the United States Environmental Protection Agency (EPA).

In contrast, high levels of copper accumulation have been observed in fish from Lake Geriyo, Nigeria. Specifically, Cu content in the liver of Clarias, Tilapia, and Heterotis varied from 22.6 ± 7.85 to 29.87 ± 9.28 mg∙kg⁻¹, 18.08 ± 6.54 to 31.15 ± 3.40 mg∙kg⁻¹, and 4.69 ± 2.05 to 20.4 ± 4.19 mg∙kg⁻¹, respectively. Notably, there is a lack of data on the correlation between metal concentrations in tissue and environmental factors [177].

The bioaccumulation of copper in the River Nile, Egypt, showed maximum Cu values in the liver tissue of Oreochromis niloticus at 6.34 μg∙g⁻¹ dry weight. The bioaccumulation factor (BAF) varied significantly based on the season and locality, reaching values as high as 11.55–2010 L∙kg⁻¹ [178]. These values are comparable to the BAF data determined for fish in the Murucupi River, an area impacted by effluents from an alumina factory located in Barcarena in the Brazilian Amazon. In particular, for Cichla spp., BAF was as high as 1130 L∙kg⁻¹, which is approximately seven times higher than the established limit; for Eigenmannia sp., it was 2885 L∙kg⁻¹, which is fourteen times the established limit; and for Angelfish, it was 1640 L∙kg⁻¹, which is eight times the established limit [179]. Additionally, these BAF values are considerably lower than those observed in the muscle tissue of Anabus testudineus, which has a BAF value of approximately 1.7 L∙kg⁻¹ [180].

It is worth mentioning that some organisms, such as mollusks and frogs, can accumulate high levels of copper without experiencing significant physiological disturbances. Turtles, being long-lived animals, have the capacity to accumulate higher levels of metals in various body tissues than in water. Consequently, they are considered appropriate indicators of chemical contamination in aquatic environments. For example, sea turtles (Chelonia mydas) stranded along the Potiguar Basin in northeastern Brazil showed copper (Cu) content of 32.737 ± 28.866 μg∙g⁻¹ wet weight, a level that might be related to their consumption of algae, the main food source for C. mydas [181]. Similar to turtles, frogs tend to exhibit Cu levels in the liver that are approximately 5–50 times higher than those found in other species, such as mollusks and fish. In frogs from rural and urban ponds in western Ukraine, the copper content varied from 39.5 ± 4.5 to 164.1 ± 14.4 μg Cu∙g⁻¹ dry weight (DW). These levels were observed to change more significantly with the seasons than with the location. Compared to zinc, cadmium, and iron, the bioavailability of Cu was approximately 1000 times higher. However, a comparison of the metal content in the frog’s liver with its metallothioneins (MTs) reveals that these proteins do not play a significant in buffering Cu in tissues. Cu was predominantly distributed in thermolabile compounds, with only 4% bound to MTs [6,182].

Mollusks, including the invasive species Dreissena polymorpha (zebra mussel), are also capable of accumulating significant amounts of copper. In the Upper St. Lawrence River, copper concentrations in D. polymorpha ranged from 30 to 60 μg∙g⁻¹ DW [183]. In contrast, lower but still comparable copper content was found in another bivalve species, Anodonta cygnea, from regions with low pollution in western Ukraine, where Cu levels ranged from 7.5 to 27.5 μg∙g⁻¹ DW. Significantly lower Cu concentrations were recorded in individuals from rural sites, particularly in the gills in autumn [34]. These mussels’ tissue concentrations tend to correlate with water concentrations, providing a sort of temporal and spatial average of contamination. Unlike point water samples, which require a high number of samples to account for all contaminant entry points, tissue concentrations in these indicator organisms reflect integrated exposure over time and across different locations. This makes them highly effective for identifying pollution trends and changes in pollutant levels.

Excessive accumulation of copper in various organs of aquatic biota, particularly in fish, which are more vulnerable to its effects than other aquatic animals, can significantly impact physiological functions [184,185]. This includes inhibited growth and development, reduced swimming activity, impaired respiration, cardiovascular collapse, and alterations in cell permeability and Na⁺/K⁺-ATPase enzyme activity. Copper exposure can also lead to metabolic disorders and an inadequate energy supply within the body. Additionally, copper may induce oxidative damage, altering the redox state of cells, leading to protein and lipid peroxidation, DNA damage, and ultimately, reduced immune functions [100,102,186].

In conclusion, the accumulation of copper in freshwater organisms varies significantly based on a multitude of factors including seasonal fluctuations, species-specific physiology, and environmental conditions. While some organisms like mollusks, frogs, and turtles have developed mechanisms to accumulate and tolerate high levels of copper, others such as fish remain highly susceptible to copper toxicity. Studies highlight that living organisms possess complex mechanisms to regulate essential metals like copper, and therefore, assessing the biorisk of metal pollution remains challenging. Seasonal variations and developmental stages are significant influencing factors. In atypical waters with pH <5.5 or >9, neither the multiple-linear regression models nor biotic ligand models predictions were reliable, suggesting that site-specific testing would be needed to determine reliable Cu criteria for such settings [187].

3.4. Bioremediation of Copper Pollution

Copper removal from the water column in both natural and artificial water reservoirs relies on natural processes involving vegetation, sediments, and microorganisms. Four primary mechanisms influence the rate and extent of metal removal in such environments: (1) adsorption to fine-grained matter and sediments; (2) precipitation as insoluble minerals; (3) geochemical cycles involving microorganisms and plants; and (4) deposition of suspended solids. While these mechanisms, through metal accumulation or geochemical precipitation, can reduce the concentration of metal ions in the water, they do not always suffice to meet regulatory standards for permissible values [188,189]. To address this issue, bioremediation can be employed, using bacteria to effectively lower metal concentrations in water through mechanisms such as sorption and biofilm formation.

The biological removal of Cu(II) in aquatic environments primarily involves four key processes: bioaccumulation, biosorption, biomineralization, and phytoremediation. Among these, the removal of Cu(II) by microorganisms stands out due to two specific mechanisms. One involves a microorganism resistance gene that enables the organisms to survive and grow in the presence of Cu(II), thereby allowing Cu(II) to be accumulated within their cells through cell membranes. The other mechanism is based on the adsorption of Cu(II) to microorganisms through both physical and chemical actions, facilitated by the secretion of extracellular polymeric substances and other compounds with adsorption capacity (Figure 1) [190,191].

Bioaccumulation involves the uptake of contaminants based on the metabolic activity of microorganisms. This process depends on microorganisms absorbing contaminants at a faster rate than they can excrete them [192]. The efficiency of bioaccumulation is directly related to the concentration of pollutants that microorganisms can accumulate. For example, Amycolatosis tucumanensis has been shown to accumulate up to 25 mg∙g⁻¹ DW of copper, with 60% of it located intracellularly [193].

Another critical factor in bioaccumulation is the capacity of low molecular weight cysteine-rich metal-buffering proteins known as metallothioneins. For instance, CUP1, which encodes a metallothionein, contributes to copper stress tolerance in Saccharomyces cerevisiae [194]. Additionally, expressing cadmium-binding metallothioneins on the surface of bacterial cells can enhance cadmium stress tolerance and the microbial bioremediation capacity of Escherichia coli [195]. Unfortunately, most research and attempts to implement bacteria for bioremediation have focused on cadmium and lead [195,196], often overlooking copper [197].

There is also a non-metabolic method of metal uptake that involves physicochemical interactions between metal ions and functional groups on the bacterial surface. These interactions include chemical processes such as ion exchange, physical interactions like electrostatic or van der Waals forces, complexation, diffusion, surface adsorption, and precipitation. In environments contaminated with heavy metals, bacteria may respond by activating CPx-type ATPases in some strains, releasing extracellular polymers with a high affinity for copper, and producing proteins capable of chelating metal ions [198,199,200].

The removal of toxic metals from aquatic media through biosorption utilizing metabolically inactive abiotic biomass, originating from both plant and microbial sources, represents a promising technology for contaminant elimination by forming metal complexes. The primary mechanism for metal biosorption, including Cu²⁺, involves ion exchange between the monovalent metals, among them Na⁺ and K⁺, present in macrophyte biomass as counterions and the toxic metal ions absorbed from the water [190]. Bacterial biomasses serve as effective biosorbents for bioremediation of metals in polymetallic conditions due to their high surface-to-volume ratio. The biosorption process, although reversible and rapid, depends on several environmental factors such as pH, ionic strength, metal concentration, and the level of dissolved organic matter. The functional groups present on microbial cell walls, composed of polysaccharides, proteins, and lipids, offer multiple sites for interaction with contaminants [190,201,202]. For instance, hydroxyl, carboxyl, amine, and sulfur groups can bind to Cu(II), resulting in organic–metal precipitates adsorbed onto microbial cells. Due to their small size, microorganisms exhibit a large specific surface area, enhancing toxic metal adsorption onto cellular polymers [202].

Biosorption efficacy varies with different types of biomass. For example, the biomass of Myrica esculenta has demonstrated a maximum observed adsorption capacity (Qmax) of 39.37 mg∙g⁻¹ for Cu(II) ions. A Fourier-transform infrared spectroscopy analysis identified functional groups such as hydroxyl, amine, carbonyl, and carboxylic groups, which interact with metal ions in synthetic wastewater. Thermodynamic analyses reveal that biosorption on Myrica esculenta biomass occurs spontaneously (ΔG° < 0), is endothermic (ΔH > 0), and feasible (ΔS > 0) under the examined conditions [203]. Pseudomonas putida CZ1 has shown remarkable tolerance and efficiency in removing toxic metals, particularly copper and zinc, from aqueous solutions, aligning well with the Langmuir isotherm model. Living cells of P. putida CZ1 exhibit significantly higher metal-binding capacity compared to nonliving cells, indicating potential for bioremediation applications [204]. Similarly, Michelia figo biomass has shown high removal efficiencies for copper, lead, and cadmium from mixed metal solutions, with a maximum removal efficiency for Cu(II), reaching 94.12% at conditions characterized by pH 5, an initial metal concentration of 0.157 mmol∙L⁻¹, and a biomass dosage of 10 g∙L⁻¹ [205].

The microbially induced carbonate precipitation process (MICP) belongs as well to the popular alternative for purifying metal-containing water. This process involves the activity of microorganisms that hydrolyze urea (in a reaction catalyzed by urease), resulting in the formation of carbonate and ammonium ions. Ammonium ions increase the pH of the environment, which in turn promotes the binding of carbonate to calcium and ultimately the formation of calcium carbonate deposits. Importantly, this mechanism is applicable to metals with a valence number of two, such as copper (II) [206,207]. Although most of the studies are related to bioremediation of contaminated soils, they are based on analyses conducted in aqueous solutions, hence these studies are useful for removing metals from aqueous solutions [208]. Achal et al. (2011) demonstrated that MICP-based bioremediation by Kocuria flava is an environmentally friendly technology for the remediation of copper-contaminated sites. They investigated the ability of calcifying bacteria strain CR1, to bioremediate copper based on microbial calcite precipitation. K. flava CR1 is able to remove 97% of copper when the initial Cu concentration was 1000 mg L⁻¹, while producing a significant amount of urease (472 U mL⁻¹), an enzyme that leads to calcite precipitation. Additionally, in the presence of urea, K. flava is able to remove 95% of Cu within 120 h, while in the absence of urea, it removes only 68%. This is evidence of the ureolytic nature of the metabolism used by the bacterium, which is further confirmed by the optimal pH of 8 for complete copper removal [209]. Kang et al. (2016) studied the synergistic effect of bacterial mixtures on the bioremediation of Pb, Cd, and Cu mixture. Four bacterial strains—Viridibacillus arenosi B-21, Sporosarcina soli B-22, Enterobacter cloacae KJ-46, and E. cloacae KJ-47—were isolated and identified from the bacterial mixtures, which showed effective microbiologically induced calcite precipitation (MICP). By modifying the bacterial mixtures, it was possible to obtain a remediation effect after 48 h of 98.3% for Pb, 85.4% for Cd, and 5.6% for Cu. Despite the low result obtained for copper, the study confirmed that, compared to single-strain cultures, bacterial mixtures showed greater resistance and effectiveness in the remediation of heavy metals, which suggests the need for the use of bacterial mixtures in the bioremediation of heavy metals from contaminated environments [210]. The great potential of bioremediation of metals, the biomineralization process, as the immobilization of toxic metal is also confirmed by the research of Qiao et al. (2021) [211]. The researchers determined the MICP efficiency for several common heavy metals (Cu, Zn, Ni, Cd) along with their precipitation patterns. Based on the urease activity and precipitation ability, Sporosarcina kp-4 and kp-22 were selected for the study. The Cd removal mechanism was the formation of cadmium carbonate induced by bacterial activity, while the removal of Cu was dependent on the increase in pH generated by the same process. The removal of Zn and Ni by precipitation was more complex. The removal rates of Cu, Zn, Ni, and Cd at the initial concentration of all metals, 160 mg∙L⁻¹, reached 75%, 98%, 59%, and 96%, respectively, within 2 h. The minimum inhibitory concentration (MIC) values showed that the toxicity of these metals to MICP bacteria was as follows: Cd > Zn > Ni > Cu, confirming that urease-producing bacteria can co-precipitate many heavy metals even without the ability to tolerate them [211].

Various studies have examined the capacity of different aquatic plants to bioaccumulate and remove Cu from contaminated water. Research involving common wetland plants such as duckweed and tape grass demonstrated that duckweed outperformed tape grass at low Cu concentrations. It was observed that leaching from decaying plant tissues did not adversely affect pond snail (Physa acuta) survival at low Cu levels, though toxicity was evident at higher concentrations, especially post-simulated freeze–thaw cycles [212].

Another study investigated eight different aquatic plant species, revealing that Eichhornia crassipes and Pistia stratiotes had superior bioaccumulation capabilities, whereas Juncus effusus, Sagittaria sagittifolia, and Echinodorus major had weaker results. Pistia stratiotes adheres to a predominantly first-order elimination kinetics. However, at higher Cu concentrations, the removal efficiency decreased [213]. Copper distribution in plant tissues varied by species, with Juncus effusus accumulating most Cu in shoot tissues, while Sagittaria sagittifolia and Acorus calamus stored Cu primarily in their roots. Notably, high lignin content in plants correlated with increased Cu accumulation [214].

Additionally, research on Limnocharis flava demonstrated its effectiveness in Cu absorption across different water types, with the highest efficiency observed in distilled water. Conversely, Hydrilla verticillata did not show measurable Cu absorption under similar conditions [215].

To sum up, bioremediation techniques for Cu (II) removal using microorganisms, fungi, and plants remain environmentally friendly approaches for treating toxic metal contamination with minimal secondary pollution. However, the bioremediator’s bioaccumulation and biosorption capacity can be influenced by the physicochemical properties of water, cell structure, and value of expression of specific metal-binding proteins. Additionally, biological methods may incur high additional costs related to environmental regulation (e.g., pH, temperature) and nutrient supplementation necessary to maintain the biological process effectively.

4. Conclusions and Perspectives

The complex issue of water pollution due to transition metals like copper and zinc has broad implications for ecosystem health, biodiversity, and human well-being. Industrialization, agriculture, and inadequate waste management are significant sources of copper and zinc pollution in water bodies. Regions like Europe, despite having strict regulations, still exhibit considerable metal concentrations in areas of intense agricultural and industrial activities. Contrarily, developing regions often show higher pollution levels due to shortcoming environmental controls.

Zinc demonstrates significant mobility due to its high solubility in oxidizing conditions and its high affinity for retention in the solid phase. Field studies suggest zinc has lower sorption capacity compared to copper, largely due to the formation of metallo-organic complexes and changes in ionic strength. In ecosystems, the specific fractions of zinc, such as those in exchangeable, reducible, and oxidizable forms, indicate its potential bioavailability and ecological risks. The environmental behavior of zinc and copper varies significantly with pH levels, where higher pH enhances the stability of zinc–natural-organic-matter complexes, reducing bioavailability.

Zinc and copper, while essential, can also be harmful due to their complex ecotoxicological dynamics. Aquatic animals have varying needs and tolerances for these metals, leading to inconsistencies between the uptake and accumulation of copper and zinc in organisms compared to their acute or chronic exposure to pollutants. Depending on their metabolic requirements, certain organisms, such as fish for copper, serve as more reliable indicators of zinc and copper pollution due to a stronger correlation between their tissue concentrations and the environmental concentrations of these metals. To develop predictive models of bioaccumulation and toxicity, it is crucial to understand the parameters that influence these correlations. These metals, at elevated concentrations, may destabilize cellular processes by binding to targeting biomolecules, leading to toxicity in aquatic animals. However, animals use different mechanisms like metallothionein regulation, mucus secretion, and lysosomal storage to counteract metal stress, illustrating both resilience and vulnerability. Despite these mechanisms, high metal concentrations induce oxidative stress, metabolic disturbances, and immune responses, suggesting significant ecotoxicological risks to aquatic life, primarily fish.

To tackle zinc and copper pollution effectively, it is essential to employ bioremediation strategies. Natural processes such as adsorption, precipitation, and geochemical cycling in wetlands and natural reservoirs are important for metal removal. Nonetheless, these natural processes need to be complemented by bioremediation techniques, among them bioaccumulation and biosorption, to achieve regulatory compliance.

Future research should concentrate on several aspects to address zinc and copper pollution effectively: (i) Innovative Biotechnological Solutions: Investigate genetic modifications in plants and microbes to enhance their capabilities for metal accumulation and detoxification. This line of research aims to develop organisms that can more effectively sequester and neutralize harmful metals, thereby improving the efficiency of bioremediation strategies; (ii) Emission Reduction Strategies: Explore promising methods to decrease the emission of copper and zinc into the environment. This could involve advancements in industrial processes, the development of eco-friendly materials, and stricter regulatory measures; (iii) Advanced Modeling Techniques: Utilize machine learning in combination with hydrochemical data to forecast pollution trends and evaluate the outcomes of remediation efforts.