Copper as a Complex Indicator of the Status of the Marine Environment Concerning Climate Change

Abstract

Studies covering key elements of the marine ecosystem based on current and long-term data have made it possible to assess both the current situation in terms of copper concentrations in commercially used fish and benthic plants and in surface bottom sediments, as well as enabled the analysis of the temporal variability of copper levels in relation to changes in its inflow to the southern Baltic Sea. By applying the threshold values, determining the boundary between good and not good status of the marine environment, set in this study, it was found that good environmental status has been achieved in the case of Cu in seawater and plants and has not been achieved in the case of sediments and fish for consumption. The study showed that climate change, the main feature of which is an increase in seawater temperature, significantly impacts the distribution and levels of copper in individual elements of the marine environment. It influences the vegetative season length and bioaccumulation efficiency and is of key importance for copper toxicity.

1. Introduction

The average copper content in the Earth’s crust is 55 ppm, of which 87 ppm is found in the basaltic rocks that are part of the igneous rocks, and 45 ppm is found in the shales that are part of the sedimentary rock [1]. At the same time, due to its wide industrial applications, copper concentrations in the environment can be much higher due to anthropogenic activities, which can lead to toxic levels in organisms. Despite a slight downward trend of total annual copper emissions from anthropogenic sources of HELCOM countries from 1990 to 2019 [2], significant amounts of copper are still discharged into the Baltic Sea. The main source of copper reaching the Baltic Sea is the riverine input, which accounts for 90.7%, the share of atmospheric deposition is 7.8%, and releases from direct point sources account for 1.5% [3]. The average copper load introduced by river waters in 2018–2021 amounted to 987 tonnes/year, with 400 tonnes/year on average in 2018 and 2019 and an increase to 1105 tonnes/year in 2020 and 855 tonnes/year in 2021. The average inflow from the Polish area in this period amounted to 59 tonnes/year, 6% of the total riverine input. Atmospheric copper deposition remained relatively stable in 2018–2021 at 85 tonnes/year [3].

Copper is considered to be very toxic [4,5], but at the same time, copper is a crucial micronutrient for organisms. Copper is one of the essential trace metals in the human organism, and third in abundance after iron and zinc [6]. The overall concentration of the element in the body is approximately 100 mg, with the greatest proportion found in the liver, accounting for 10% of the total amount [7]. The oxidised Cu(II) and reduced Cu(I) forms are pivotal for cell physiology since they act as a catalytic cofactor for enzymes and participate in mitochondrial respiration, iron absorption, antioxidant activities and elastin cross-linking. Several necessary cellular enzymes, including cytochrome-c oxidase, copper/zinc superoxide dismutase (Cu/Zn-SOD), metallothioneins, ceruloplasmin, hephaestin, cartilage matrix glycoprotein (CMGP), and lysyl oxidase, need copper for their function. The impact of Cu is easily recognised in the global antioxidant response since the Cu/Zn superoxide dismutase 1 (SOD1) is a main free-radical scavenger within the cell, and its mutations cause amino acid, lipid, and nucleic-acid damage. Mutant versions of SOD1 increase oxidative stress, contributing to the development of numerous diseases, for instance, hepatocellular carcinoma, macular degeneration, and amyotrophic lateral sclerosis. Viewed collectively, Cu-dependent functions of cytochrome-c oxidase, SOD1, and glutathione indicate the importance of copper as a static cofactor in energy production and cellular redox metabolism [8].

Another vital aspect of the role of copper is connective and bone tissue formation due to the cuproenzyme lysyl oxidase (LOX), which facilitates cross-linking of collagen and elastin fibres. Moreover, interactions of Cu and Fe in biological systems via copper-dependent ferroxidases such as ceruloplasmin and hephaestin mean that copper scarcity precipitates limited iron absorption, thus causing defective erythropoiesis and microcytic anaemia due to iron depletion. Another Cu-containing enzyme, cytochrome-c oxidase (COX), is necessary for synthesising sphingolipids, which constitute a part of the myelin sheath. Therefore, the deficit in copper might impair central nervous system development and post-development functioning. Nevertheless, redundant Cu can be both cytotoxic and genotoxic. Similarly to iron, copper is involved in reactions producing highly reactive oxygen species (ROS), thus leading to direct oxidations of proteins, lipid peroxidation in membranes, and damage to DNA and RNA molecules [9].

In the presence of redox agents, such as ascorbic acid or glutathione, copper catalyses the synthesis of highly reactive hydroxyl radicals from hydrogen peroxide in the Haber–Weiss cycle [10]. These radicals initiate oxidative damage in various ways. Additionally, Cu ions accelerate lipid peroxidation by breaking lipid hydroperoxides in a process similar to the Fenton reaction, resulting in aloxyl and peroxyl radicals, which propagates the chain reaction [11]. Apart from the generation and action of ROS, copper demonstrates its toxicity by the displacement of other metal cofactors from their naturally existing ligands [8]. Copper genotoxicity at the chromosomal level can be measured with a cytogenetic tool, the Cytokinesis-Block Micronucleus Cytome (CBMN Cyt) assay, since it analyses the incidence of micronuclei (biomarker of breakage or loss of chromosomes), nucleoplasmic bridges (biomarker of DNA strand breakage, misrepair, and telomere end fusion), and nuclear buds (biomarker of DNA repair complexes elimination and gene amplification). This bioassay also allows assessing the cytotoxicity of copper via nuclear division index (NDI) and a number of apoptotic or necrotic cells. Applying this method for WIL2-NS human B lymphoblastoid cells shows considerable cytogenotoxic damage is observed with copper at the concentration of 10–100 µM, and total cytotoxicity is noticed at 1000 µM [12].

According to the research with different copper extracts administered to Chinese hamster ovary cells (CHO-K1), the reduction of mitochondrial activity caused by Cu ions is recognised at ≥7.42 mg/L concentrations, decrease in cell viability at 10.85 mg/L, as well as copper-induced DNA damage at 5.67–7.42 mg/L concentration range [13]. Having numerous proofs of Cu-induced toxicity in human organisms, safe upper levels of copper intake per day were proposed: 10–12 mg/day by the International Program on Chemical Safety (1998) and the World Health Organization (1996), 10 mg/day for adults by Institute of Medicine (US) Panel on Micronutrients (2001), and 0.07 mg/kg of body weight per day for adults by The European Food Safety Authority (2023) [14].

Copper as a microelement also plays a key role in marine organisms, both fauna and flora. In algae, it is crucial for ensuring electron transport in photosynthesis and the proper functioning of enzymatic systems [15]. At the same time, as in humans, it may have toxic effects, i.e., inhibit photosynthesis, disrupt electron transport, reduce pigment concentration, affect the permeability of plasma membrane, inhibit nitrate uptake, restrict growth, and affect the distribution of other compounds (e.g., lipids, proteins, and sterols) [15] w where its concentration exceeds acceptable levels. Copper concentration in water above 0.1 mg/L slows down the growth of macrophytes [16]. At a copper concentration of 0.45 mg/L, the development of humpback duckweed is inhibited by approximately 50% [17]. At the same time, algae use protective mechanisms involving exclusion by producing metal-binding compounds [18] or intracellular detoxification [19].

Copper toxicity to aquatic organisms depends on its “bioavailability,” which determines its transfer to receptors like gills and olfactory neurons [20]. In fish, copper effects are “acute” (lethal) or “chronic” (sublethal), impacting growth, immune response, reproduction, and survival [21,22]. Copper adversely affects gills, olfactory receptors, and lateral line cilia, reducing disease resistance, disrupting migration, altering swimming, causing oxidative damage, impairing respiration, and affecting osmoregulation and organ structure. It also changes behaviour, blood chemistry, enzyme activities, corticosteroid metabolism, and gene expression [23,24].

Taking into account the potential risks associated with the presence of copper in the marine environment, both related to the harmful effects on marine organisms and humans due to fish consumption, studies were carried out to determine the actual levels of Cu in selected elements of the southern Baltic ecosystem, including seawater as the primary source of Cu, bottom sediments as the final reservoir, and selected species of fish and algae. The complex indicator described as copper levels in various (living and non-living) elements was proposed to assess the state of the environment. It covers multiple aspects of the entire ecosystem (distribution in the water column as a result of the introduction of Cu to the marine environment, bioaccumulation in organisms (fish and plants), and accumulation in bottom sediments. The research was complemented by fish blood analysis using a micronucleus test to indicate genotoxic effects. An analysis of the risks associated with fish consumption in relation to the acceptable levels of Cu was carried out. Taking into account the current physicochemical characteristics of the southern Baltic area and potential changes resulting from climate change, the levels of Cu in seawater corresponding to toxic conditions were determined in this area. Quality standards have been set for macrobenthic plants and bottom sediments. For fish, the threshold value is based on acceptable levels of copper in fish for consumption.

2. Materials and Methods

2.1. Study Area and Matrices

The study area covered the southern Baltic Sea, including the Gulf of Gdańsk with discriminated Puck Bay, the Gdańsk Deep, the eastern Gotland Basin, the Bornholm Basin, and Pomeranian Bay (Figure 1). The studies were also conducted in the Szczecin Lagoon and the Vistula Lagoon. In the area of research, the following fisheries were also distinguished: Kołobrzesko-Darłowowskie (LKOL) and Władysławowskie (LWLA). Studies of copper levels included key elements of the marine environment: fish (muscles), macrobenthic plants, and bottom sediments. Studies using the micronucleus test were carried out in fish blood samples.

2.2. Sampling

2.2.1. Fish for Copper Analysis

The study covered three fish species: herring (Clupea harengus) collected in 1994–2022 from the Kołobrzesko-Darłowskie fishery and in 1998–2022 from the Władysławowskie fisheries, flounder (Platychtys flesus) collected in 2012–2022 from the Pomeranian Bay and in 2016–2022 from the Gulf of Gdańsk, and perch (Perca fluviatilis) collected in 2004–2022 from the Szczecin Lagoon and in 2014–2022 from the Vistula Lagoon (Figure 1). Each time, 10 to 15 individuals (females) were taken for the study. Muscles were taken from each specimen, and Cu concentrations were tested. Since 2014, the collected individuals have undergone ichthyological analysis to determine the following biometric parameters: length [cm], weight [g], stage of gonadal development, age based on otolith analysis, and sex.

2.2.2. Fish Blood for Micronucleus Test Analysis

Studies were carried out using the micronucleus test (MN) involving herring in 2014–2022 and perch in 2018–2022 to determine the impact of hazardous substances on marine organisms. The micronucleus test is the most commonly used test to assess cytogenetic damage at the cellular level caused by exposure to hazardous substances [25]. Blood samples were collected from commercially caught fish in the southern Baltic region: Baltic herring (Clupea harengus) and perch (Perca fluviatilis) from areas most similar to the fish fisheries used for Cu analyses in muscles (Figure 1). Blood collected from the caudal fin was applied to slides, stained with Giemsa reagent (1:9) and microscopically analysed using the brightfield technique at 1000× magnification. The observation consisted of counting abnormalities (micronucleated changes) occurring within the cell according to established criteria [25]. The number of counted micronuclei converted to 1000 erythrocytes is a parameter that measures the harmfulness of the impact of hazardous substances on the organism [26].

2.2.3. Sediment

Sediments for copper studies were collected in six locations, including four open sea areas of silty sediment formation characterised by the presence of a dominant fraction below 63 μm and a significant content of organic matter, which guaranteed copper detection (stations P1, P140, P5, P39—Figure 1) and in two lagoon areas: the Vistula Lagoon (KW) and the Szczecin Lagoon (GJ), where silty sediments also occur. Sediments were collected using a core probe, and the sediment cores were divided into layers 2 cm thick to a depth of 10 cm. Then, 2 cm thick layers were collected, discarding the 5 cm layers. Finally, copper analyses were carried out in layers of 0–2 cm, 2–4 cm, 4–6 cm, 6–8 cm, 8–10 cm, 15–17 cm, 22–24 cm, 29–31 cm, and 36–38 cm in the case of open sea areas. In the case where the length of the core allowed it, samples were also taken from deeper layers: 43–45 cm, 50–52 cm, 57–59 cm, 61–63 cm, and 68–70 cm. The samples were taken into plastic vessels, frozen, and dried by freeze-drying before analysis. Sediment samples were collected in 2012–2021 with different temporal resolution depending on the location. For the P5 station, samples taken from 2002–2018 were used in the analyses to check the reliability of the results.

2.2.4. Macrobenthic Plants

For copper analysis in a macrobenthic plant, samples were taken in five locations: in the Gulf of Gdańsk (KO) in 2008–2010 and 2013–2022, in the Bay of Puck (JK), in the coastal area (RO) and Slupsk Bank (ŁS) in 2010, 2013–2022, and the Wolin National Park (WP) in 2019–2021 (Figure 1). Algae species representing red algae and green algae, vascular plant species and one species representing charophytes were collected for the study. Sampling was carried out during the peak of plant growth in June and during the slow growth of macroalgae in September. In the case of the Orłowski Cliff and Jama Kuźnicka transects, samples were taken from a depth of 1 m to the maximum depth of plant occurrence. In the Slupsk Bank and the Rowy Boulder, sampling was carried out at previously designated points due to the lack of constant depth change. Plant material was collected from 0.25 m × 0.25 m, limited by the frame, and randomly placed four times at each depth. In the case of hard bottom, organisms were taken from the surface of stones. The collected material was stored in a frozen state at −18 °C until analysis [27].

2.3. Analysis

2.3.1. Copper Analysis

• Fish

Prior to Cu analysis, the collected muscle tissue was mineralised in nitric acid using Milestone’s Ultra Wave microwave mineralizer. In the obtained mineralizates, the copper concentration was determined using the flame atomic absorption (AAS-FL) and electrothermal (AAS-GF) methods using spectrometers from Thermo Scientific. In each series, standard solutions and a blank sample were analysed. To ensure the quality of the measurements, metal determinations were performed in parallel with the Cu determinations in fish in six samples of Dolt-5 certified material. The accuracy of the method was calculated as the average recovery percentage in the five samples at 91%. The repeatability of the method was determined as the relative standard deviation for the six parallel Dolt-5 material determinations, which was at the level of 2%. The limit of copper quantification in fish was 0.03 mg kg⁻¹ wet weight.

• Plants

Before the copper analysis, the collected plant material was washed 4–5 times in trays filled with distilled water with a volume of not less than two litres. The material was analysed taxonomically according to the guidelines [27]. Individual species occurring in the samples were identified, separated, placed in plastic string bags, and frozen. Taxonomic analyses and biomass determination were carried out. For copper determination, the selected plant species were freeze-dried. The dry biomass of each species was determined gravimetrically. Then, samples were ashed in platinum vessels at 450 °C and mineralised in ultra-pure nitric acid in a Milestone Ultra Wave mineraliser. This allowed the process to run at circa 300 °C and pressure up to 200 bars. The glassware used for the sample preparation was cleaned with ultra-pure nitric acid with a concentration of approximately 10% for three days. Before direct use, the glass was rinsed several times with deionised water. Copper content in plant samples was determined using a flame atomic absorption spectrometer—Scientific ICE 3300. To ensure the quality of measurements, metal determinations were performed in parallel with Cu determinations in phytobentos in six samples of MPH-2-certified material. The accuracy of the method was calculated as the average recovery percentage in the five samples at 96%. The repeatability of the method was determined as the relative standard deviation for six parallel MPH-2 material determinations of 4%. The limit of copper quantification in phytobentos was 0.035 mg kg⁻¹ dry weight.

• Sediment

Before copper analysis, the dried sediment samples were homogenised. Copper concentrations were measured in the sediment mineralizates, obtained by treating the sediment samples (ca. 1.5 g) with concentrated acids HNO₃ and HF. The mineralisation was carried out in Teflon vessels at elevated temperatures. The metal concentrations were measured using atomic absorption spectrometry (AAS). To ensure the quality of measurements, Cu analyses were performed in six samples of certified MESS-4 material. The accuracy of the method was calculated as the average percentage recovery in the samples at the level of 93%. The repeatability of the method was determined as the relative standard deviation for six parallel determinations of MESS-4 material at the level of 7%. The limit of quantification of copper in sediments was at the level of 0.035 mg kg⁻¹ dry weight.

2.3.2. Sediment Dating

To trace historical changes in copper concentrations in bottom sediments, the results of dating bottom sediments carried out in 2009 and 2020 were used [28,29]. Samples for dating were taken in the exact locations for Cu analysis. The dating was carried out using an isotope method based on measurements by gamma spectrometry of radioactive activity of ²¹⁰Pb and ¹³⁷Cs, with caesium being used to verify the correctness of the results of modelling the age of sediments. Dating was carried out using the following models: the Constant Rate of Supply (RSC) model (based on the assumption that the supply of ²¹⁰Pb to the sea surface is constant, while the sedimentation rate might vary) and the Constant Flux Constant Sedimentation Rate (CF:CS) model (assumes a constant dry-mass sedimentation rate). The ²¹⁰Pb dating method allows determining the age of sediments up to approximately 150 years based on measurement data. Using the dependence of sediment age on depth described by the second-degree polynomial equation in individual locations, extrapolation beyond the measurement range was carried out, and the age of deeper sediment layers was determined.

2.3.3. Processing Results

Statistical analysis and graphic representation of the results were performed using STATISTICA 13.3 (StatSoft 2023/2024). The analysed data were characterised by non-normal distribution (Shapiro–Wilk test p < 0.05). The ANOVA Kruskal–Wallis tests were used to determine the significance of differences between the Cu concentration and biometric parameters. The relationships between the analysed variables were determined based on Spearman’s coefficient, with a confidence interval of at least 95%. p < 0.05 was regarded as a statistically significant difference.

The estimated daily intake (EDI) expressed in mg kg⁻¹ body-weight day⁻¹ of the Cu was calculated by using the below equation as reported by [30]:

EDI = (C_element × D_{food intake})/BW

(1)

where C_element: is the concentration of HM in the muscle tissue of fish (as mg kg⁻¹ wet weight), D_{food intake}: is the mean food (fish) consumption daily (g/person/day), that is 2.88 kg per person per year in Poland in 2022 and BW is the mean body weight (70 kg for adults).

3. Results and Discussion

3.1. Copper in Fish and Micronucleus Test Results

Fish, typically at the top of the trophic level in aquatic habitats, absorb heavy metals (HMs) from their environment due to factors like species characteristics, exposure duration, element concentration, and water parameters [31,32]. This study found that commercial fish in the southern Baltic Sea contain varying concentrations of copper (Cu), with accumulation levels differing among species. The mean copper concentration during the study period in herring muscles in the Kołobrzesko-Darłowskie fishery was 3.32 mg/kg mm. The minimum value, equal to 1.23 mg/kg mm, was determined in 2020, while the maximum value in 1995 was equal to 6.69 mg/kg ww.

In the case of the Władysławowskie fishery, the mean Cu value was 3.40 mg/kg mm. The minimum value of 1.61 mg/kg mm was determined in 2007, with a maximum value of 5.37 mg/kg ww in 1994. The mean copper concentration during the study period in flounder in the Pomeranian Bay area was 15.34 mg/kg mm. The minimum value, equal to 4.01 mg/kg mm, was determined in 2013, and the maximum value in 2020, equal to 28.59 mg/kg ww. In the case of the Gulf of Gdansk, the mean value of Cu was 4.59 mg/kg mm. The minimum value of 1.71 mg/kg mm was determined in 2022, with a maximum value of 14.89 mg/kg ww in 2017. The mean copper concentration during the study period in perch muscles in the Szczecin Lagoon area was 4.59 mg/kg mm. The minimum value, equal to 1.57 mg/kg mm, was determined in 2006, and the maximum value in 2014, equal to 10.47 mg/kg ww. In the case of the Vistula Lagoon, the mean Cu value was 4.49 mg/kg mm. The minimum value, equal to 1.67 mg/kg mm, was determined in 2017 and 2022, while the maximum value in 2019, equal to 12.00 mg/kg ww.

The highest mean concentration of copper in muscles during the study period was determined in flounder: 12.0 mg kg⁻¹ ww and was statistically significantly higher (Kruskal–Wallis test, p < 0.05) than the average concentration of Cu in herring muscles: 3.8 mg kg⁻¹ ww and perch: 4.2 mg kg⁻¹ ww (

Table 1. Correlation of copper concentrations in individuals of three species with ichthyological parameters determined using the Spearman rank method (for values in bolded red: r-value: p < 0.05).

Species	Herring	Flounder	Perch
Mean Cu (mg kg−1 ww)	3.8	12.0	4.2
		r coefficient
length [cm]	−0.07	0.05	0.29
mass [g]	−0.32	0.10	0.30
gonads stadium	−0.54	−0.24	−0.10
age	−0.19	0.16	−0.07

). This relationship results from the life niche and food preference characteristic of a particular species. The lowest concentrations of the metal were determined in planktivorous herring (Table 1). The inversely proportional relationship between Cu concentration and the mass (r = −0.32) and age (r = −0.19) of herring may indicate the effect of metal biodilution in this species (Table 1). Metal concentrations increase with age and size when the growth of the organism is slow relative to the rate of metal accumulation [33]. In contrast, in this case, the rate of organism growth outweighed the bioaccumulation of Cu. Although in predatory perches, the average concentration of Cu in the muscles was at a similar level and did not differ statistically significantly from that of herring, an inverse relationship was observed. The copper concentration was directly proportionally correlated with the mass (r = 0.29) and length (r = 0.30) of P. fluviatilis individuals (Table 1). This indicates the process of bioaccumulation of Cu in the muscles of perch during the growth process, which is associated with a change in diet in the size classes of this species: from zooplanktonic juveniles to predatory (adult) individuals feeding on small fish [34]. Statistical analysis showed bioaccumulation of copper in the muscles of benthivorous flounder, reaching the highest concentrations in the oldest individuals, and its average concentration was three times higher compared to the other species studied (Table 1). Macrozoobenthos organisms, the main food of flatfish (including flounder), are characterised by up to twice higher concentrations of metals than sediments [35,36]. Nowadays, when the inflow of metals to the seas is lower, the remobilisation of metals from the sediment is becoming more important as a source. As a result of several biotic and abiotic processes, Cu has become remobilised into the marine environment. In this way, it can be bioaccumulated in benthic organisms and accumulated in the subsequent links of the trophic chain, thereby representing the greatest threat to marine ecosystems, including fish [37]. Hence, this could explain a significantly higher concentration of Cu in benthic flounder compared to other fish species.

In the present study, the temporal variability of Cu concentration in the muscles of the three studied fish species was also characterised, taking into account spatial changes (fishery area, Figure 2). In the case of C. harengus, a statistically significant decrease in the concentration of the metal was observed from 5.0 mg kg⁻¹ ww in 1995 to 2.8 mg kg⁻¹ ww in 2022 in the area of the Kołobrzesko-Darłowskie fishery and from 4.6 mg kg⁻¹ ww in 1998 to 2.5 mg kg⁻¹ ww in 2022 in the area of the Władysławowo fishery (Figure 2a,b). The mean concentration of Cu in herring muscles did not differ statistically significantly between fisheries (p > 0.05). The average concentration of Cu in flounder muscles was statistically significantly higher in the Pomeranian Bay region (Kruskal–Wallis test, p < 0.05), where an increase in copper concentration was observed from 8.0 mg kg⁻¹ ww in 2012 to 16.0 mg kg⁻¹ ww in 2022 compared to the Gulf of Gdańsk, where the concentration from 2016 decreased almost six times compared to 2022, reaching an average value of 2 mg kg⁻¹ ww (Figure 2c,d). In the case of perches, the statistical analysis of the time trend did not show a significant statistical change in Cu in the muscles of P. fluviatilis. However, over the years, a downward trend was observed until 2022 in the case of populations from the Vistula Lagoon and the Szczecin Lagoon (Figure 2e,f). Also, no statistically significant difference in Cu concentration was found between fisheries (p > 0.05).

The observed decrease in Cu concentrations in the muscles of herring and perch from both fisheries and flounder from the Gulf of Gdańsk is related to the reduction of Cu concentrations in the river waters of the Vistula and Oder, which are the main source of metals discharged into the southern Baltic Sea. The most significant decrease occurred until 1995, and the current concentrations are at the level of 1.45 µg L⁻¹ in Vistula and 1.85 µg L⁻¹ in Odra (Figure 3), reflected in the copper loads discharged into marine waters.

The results obtained for copper were supplemented by analyses of a micronucleus (MN) test in the blood of fish from similar fisheries, which is a measure of the genotoxicity of specific substances present in the environment [25]. Some metals can cause nuclear abnormalities. The formation of nuclear abnormalities has been reported in fish erythrocytes due to exposure to environmental and chemical contaminants of cytotoxic, genotoxic, mutagenic or carcinogenic action [38]. Higher copper concentrations in fish muscles did not correlate with an increase in the occurrence of aberrations in the blood of herring and perch (Figure 4a,b). Differences in erythrocyte micronucleus (MN) frequencies are generally linked to cell kinetics and replacement. This may explain our results, which showed that while fish exposed to heavy metals, both individually and in combination, accumulated these metals in their bodies, there was no increase in MN frequency in peripheral blood erythrocytes. It is possible that this fish species has a defense mechanism that protects against toxic substances in the early stages of exposure. Micronucleus and nuclear abnormality tests revealed that exposure to Cu and Cd did not significantly increase MN frequency in Gambusia affinis [39].

The trend of change in the occurrence of aberrations in blood was identical to the change in Cu concentration in fish muscles, where in 2022, the MN test result reached the value below the threshold value of good environmental status (GES): 0.39 (Figure 4a,b).

3.2. Copper in Macrobenthic Plants

Macrobethic plants play an important role as indicators of contamination of the marine environment with pollutants due to their ability to accumulate metals directly from the surrounding environment [40,41,42]. The concentrations observed in their tissues are proportional to those observed in water, and the response rate to environmental changes is virtually instantaneous [43]. The bioaccumulation efficiency of an element is primarily determined by its bioavailability resulting from concentrations and forms (speciation). Copper in the marine environment occurs as free copper, organically complexed and inorganically complex [15]. The most available form of copper is the ionic form, while the most common form is copper complexed by organic ligands (the content of this form can be greater than 90% depending on the availability of dissolved organic matter). The ability to accumulate copper depends on the type of ligand, which can reduce toxicity or increase it by facilitating its introduction into the tissue, as is the case with, e.g., hydrophobic compounds [15]. Copper concentrations were determined in 25 species of macrobethic plants representing green algae, red algae, brown algae, characea, and vascular plants (Figure 5). In the case of green algae, Cu concentrations ranged from 3 mg kg⁻¹ dw in Rhizoclonium riparium to 30.5 mg kg⁻¹ dw in Ulva flexuosa. It should be emphasised that the analyses of most green algae species included only a single sample during the study period. The most common species of green algae is Cladophora glomerata, for which the average concentration of Cu was 13.2 mg kg⁻¹ dw. The concentration of Cu in the only species representing brown algae (Pylaiella littoralis) was 10.7 mg kg⁻¹ dw. Similar values were specific to vascular plant tissues (6.7–19.3 mg kg⁻¹ dw). Differentiation in concentrations occurred for two species of Characea, with a significantly higher concentration in Tolypella nidifica determined in only one sample due to the rarity of this species. Tissues of all red algae species, which constituted the most numerous group of plants due to their prevalence, were characterised by higher concentrations of Cu, which ranged from 19.7 mg kg⁻¹ dw to 31.6 mg kg⁻¹ dw. The best-characterised species in terms of bioaccumulation capacity of metals and radionuclides in the southern Baltic area is Vertebrata fucoides [40,41,42], in which copper concentrations ranged from 3.8 mg kg⁻¹ dw to 83 mg kg⁻¹ dw in 183 samples.

The analysis of temporal changes in the most common and most numerous species in terms of biomass showed the lack of a clear downward trend that occurs in the case of fish or bottom sediments, which can be identified with the decline in the inflow of Cu with river waters. In 2018 and 2019, a visible increase in the average annual concentrations of Cu in red algae tissues was observed compared to the previous period (Figure 6a). A significant decrease in concentrations in four red algae species was recorded in 2022. Similar characteristics were found for vascular plants, with the highest concentrations of Cu in three of the four species studied recorded in 2019, followed by a systematic decrease (Figure 6b). Such variability is directly related to changes in seawater temperature in individual years. In the years 2008–2017, the average annual temperature of the surface layer of the southern Baltic Sea remained in the range of 10.9–12.8 °C, and in the 0–10 m layer, it was 10.8–12.0 °C [44]. The year 2018, with an average annual temperature of 17.8 °C in the surface layer and 15.3 °C in the 0–10 m layer, was the warmest year since the beginning of regular monitoring conducted in 1969 [44]. In 2019, the average annual value dropped to 13.0 °C, and in the following years, it remained at 12 °C. The strong relationship between seawater temperature and Cu levels in macrobenthic plant tissues clearly indicates an increase in the rate of Cu bioaccumulation in both algae and vascular plants resulting from temperature increases. Such a relationship, assuming an increase in temperature resulting from climate change, will have an impact on copper levels in plants, which may be reflected in the harmful effects on the plant organisms.

Climate change is causing a shorter or no icing period, which affects the circulation of metals between the overwater, sediments and benthic organisms [45]. A similar trend was observed in the case of Cu in the case of Hg, which suggests that the extension of the growing season of macrophytobenthos may be one of the factors causing an increase in metal concentrations. Previous observations regarding other heavy metals and the radioactive isotope Cs-137 [40,41] are confirmed by the results of this study showing that at lower biomass, which is the case at greater depths, the concentrations of Cu in the V. fucoides increase (Figure 7a). This applies to the average for all periods of biomass uptake and to the sum of biomass taken up in different periods from the same depths. At depths from 1 to 4 m, where the biomass reaches maximum values, the average concentration of Cu in V. fucoides is at the level of 20–25 mg kg⁻¹ dw. Below, there is a slight increase at 7 m depth and the largest at 8 m depth to 40 mg kg⁻¹ dw, where biomass size is already at the level of single grams per 1 m². Integrated results of Cu concentrations in V. fucoides from five locations confirm the increase in Cu content in tissues of V. fucoides with depth (Figure 7b). Such relationships result from the dilution effect in the case of larger biomass.

3.3. Copper in Sediments

Analysis of copper concentrations in stratified bottom sediment cores allows us to determine historical changes in copper levels in the marine environment. Copper introduced into marine waters undergoes form transformations, bioaccumulation, and sorption, which is finally deposited in bottom sediments. In the Bornholm and Gdańsk Basins, periods of limited anthropogenic impact can be distinguished, characterised by relatively constant copper concentrations (Figure 8). In the Bornholm Basin, which is outside of the direct influence of rivers, until about 1950, Cu concentrations remained in a narrow range from 35 mg kg⁻¹ dw to 39 mg kg⁻¹ dw for all years except 2018, in which concentrations in the deepest layers remained in the range of 40–42 mg kg⁻¹ dw. After 1950, there was a visible increase in concentrations related to the intensification of industrial development. Still, its signal is shifted in relation to the actual period of the Industrial Revolution. Current concentrations in the Bornholm Basin are at the level of 40–45 mg kg⁻¹ dw.

In the Gdańsk Basin, which is under the direct influence of the waters of the Vistula, an increase in copper concentrations has been observed since about 1850, corresponding to the intensification of industrial development (Figure 8). Until then, the copper level remained in the range of 35–40 mg kg⁻¹ dw, which corresponds to the Cu content in sedimentary rock shales (45 ppm) and in deep-sea sediments (30 ppm) [1]. The increase after 1850 is continuous until 2000, after which a stabilisation or slight decrease in Cu concentrations is observed, which is more significant in the Bornholm Basin. The observed trends reflect changes in Cu concentrations in river waters, which have been much lower since 2000 (Figure 4). The current concentrations in the Gdańsk Basin are at the level of 50–55 mg kg⁻¹ dw.

In the eastern Gotland Basin, Cu concentrations remain at 35–40 mg kg⁻¹ dw from 1850 to 2018. A different characteristic was noted in the sediment cores of the Vistula Lagoon, where in the years 1850–1950, the concentrations slightly decreased from 30 mg kg⁻¹ dw to 25 mg kg⁻¹ dw, after which the decrease in the level of concentrations in the bottom sediments intensified to 15 mg kg⁻¹ dw—the value observed today (Figure 8). It could relate to the limitation of copper inflow from point sources and the waters of rivers Szkarpawa, Nogat, Elbląg, and Pasłęka entering the Vistula Lagoon. The highest concentrations of copper in bottom sediments were recorded in the Szczecin Lagoon, a natural reservoir of pollution brought by the waters of the Odra (Figure 8). In the area of the Szczecin Lagoon, there are point sources like ports, and a shipping canal connecting Szczecin and Świnoujście runs through it. Until 1950, Cu concentrations did not exceed 30 mg kg⁻¹ dw, after which there was a significant increase to the maximum values of 70 mg kg⁻¹ dw around 2000, after which the copper content stabilised and decreased slightly.

3.4. Copper Toxicity and Environmental Assessment

Toxicological data and analysis of copper concentrations in various elements of the southern Baltic Sea, considering changes in the historical aspect, were used to set threshold values defining the boundaries between good and not good environmental status. This is essential for a reliable assessment of the state of the marine ecosystem in terms of copper contamination, which, under certain conditions, can pose a significant threat to marine organisms and humans. The need to define good environmental status by setting quality standards (threshold values) also results from the requirements of the legislation, in the case of the Baltic Sea as one of the EU marine areas from the provisions of the Marine Strategy Framework Directive (Directive 2008/56/EC of the European Parliament and of the Council of 17 June 2008 establishing a framework for community in the field of marine environmental policy). The study also considered the anticipated changes in the physicochemical characteristics of the waters of the southern Baltic Sea in connection with the observed climate change.

The first step was determining Cu concentrations in seawater corresponding to toxicity conditions. Models describing the dependence of acute toxicity on the parameters characterising the physicochemical conditions of marine areas (temperature—T, pH, salinity—S) were used for the calculations [4,46]. Equation (2) was used to calculate acute toxicity depending on pH, and T [4]. Acute toxicity was also determined based on single parameters: T (Equation (3)), pH (Equation (4)), and S (Equation (5)) [46].

$${Log}_{10}{T}_{acute}=-1.36+0.46\times pH-0.029\times T$$

(2)

$$Ln{T}_{acute}=6.55-0.1\times T$$

(3)

$$Ln{T}_{acute}=-8.64+1.53\times pH$$

(4)

$$Ln{T}_{acute}=-0.004\times S^2+0.13\times S+3.17$$

(5)

Studies have shown that the acute toxicity of Cu increases with temperature and decreases with increasing pH [4,44]. This means that the process of ocean acidification will also increase Cu toxicity. Still, in the conditions of the Baltic Sea, the eutrophication effect, including the intensity of blooms and the use of CO₂ in photosynthesis, causes the pH in surface waters to increase (unpublished results). The dependence of copper toxicity on salinity is more complex, as in the range of 0 to 15, the toxicity increases, while at higher salinity, it begins to decrease [46]. The calculations were carried out using data from 2008–2022 and for 2100, considering the expected change in seawater temperature [44], pH and salinity (unpublished data) resulting from climate change. For the calculations, the maximum values of the temperature in each year were taken for calculations to account for extreme conditions. For most years, the copper concentrations corresponding to acute toxicity were at the level of 100 μg L⁻¹. In 2014 and 2018, when the seawater temperature of the southern Baltic Sea exceeded 24 °C, acute toxicity could occur at a concentration of 60 μg L⁻¹ (Figure 9).

Assuming an increase in the temperature of the southern Baltic Sea by 0.6 °C per decade [44], the forecast maximum temperature will be 27 °C, and acute toxicity may occur already at concentrations below 50 μg L⁻¹. Considering pH levels, acute toxicity occurs in the range of 45 μg L⁻¹ to 80 μg L⁻¹, and assuming an increase in pH in surface layers, the toxicity limit can be shifted to approx. 100 μg L⁻¹ (Figure 9). Considering both parameters, the toxicity is determined by the range of concentrations 65–95 µg L⁻¹. Due to the lack of significant changes in the salinity of the surface waters of the southern Baltic Sea, the copper concentration corresponding to the conditions of acute toxicity is also relatively constant. Chronic toxicity values calculated based on T and pH remained in the range of 7.1 μg L⁻¹ to 13.5 μg L⁻¹ during the study period, while assuming forecasts for 2100, it may fall to 5.7 μg L⁻¹ (Figure 9). The current legislation in Poland allows copper levels up to 10 μg L⁻¹.

The current recommendation on acceptable copper levels in marine waters is based on the Voluntary Risk Assessment Report (VRAR) for copper, copper(II) sulfate pentahydrate, copper(I) oxide, copper(II) oxide, dicopper chloride trihydroxide submitted to ECHA based on industry initiative to follow the risk assessment procedures of Existing Substance Regulation (EEC) No 793/93 [47]. In the VRAR-Cu, Predicted-No-Effect-Concentrations (PNEC) were developed for terrestrial, freshwater, and marine ecosystems based on the chronic toxicity data used to prepare species sensitivity distribution (SSD) curves. PNEC-value was derived using the 50% confidence value of the SSD curve’s 5th percentile (HC5). The reported HC5 value was 5.2 μg L⁻¹ (when normalised to DOC = 2 mg L⁻¹). An assessment factor (AF) of 2 was applied to the marine environments as a precautionary approach. Finally, a PNEC_marine of 2.6 μg Cu/L was proposed and accepted by the European Commission’s Scientific Committee on Health and Environmental Risks (SCHER).

Cui and his team set water quality standards considering both the current situation and the changes in the marine environment in terms of temperature and pH projections based on climate scenarios (RCP2.6, RCP4.5, and RCP8.5) [4]. They used an analogous method by determining WQS as the 5th percentile of SSD curves based on acute and chronic toxicity data. The results showed that the short-term water quality standards values (SWQC) were 1.53, 1.41, 1.30, and 1.13 μg L⁻¹ for 2020, 2099-RCP2.6, 2099-RCP4.5, and 2099-RCP8.5, respectively [4]. The WQS value for the marine environment, similar to that determined in the VRAR-Cu, is recommended by the United States Environmental Protection Agency (USEPA)—4.8 μg L⁻¹ [48]. Based on the PNEC value determined in the VRAR-Cu without applying the assessment factor and indicated by the USEPA, it could be recommended to adopt a new value for the southern Baltic equal to 5 μg L⁻¹ (WQS).

Based on the data on the levels of Cu concentrations in coastal waters obtained as part of the State Environmental Monitoring in the years 2011–2022, no copper concentrations in seawater were found that would correspond to the toxicity conditions and remain below the recommended threshold value. The average copper concentration in coastal waters for the period 2011–2022 was 3.2 μg L⁻¹. In 2016–2017, the average annual copper values for all water bodies ranged from 1 to 4 μg L⁻¹. In 2021–2022, the highest average values were recorded in the Szczecin Lagoon 8 μg L⁻¹ and the Vistula Lagoon 5 μg L⁻¹. In 2022, in the transitional and coastal waters of the West Pomeranian coast, the annual average values of copper in water ranged from 2 to 3 μg L⁻¹, while for the eastern coast, the values were below the limit of quantification (LoQ).

Based on the newly adopted in this study threshold value for water (WQS) and the characteristics of the bioaccumulation capacity of the studied macrobenthic plant species, the value of the environmental quality standard for plants (PQS) (

Table 2. Mean copper concentrations in macrobenthic plants, bioconcentration factors, and plant quality standards.

		Mean (Min–Max) Cu Concentration (mg kg−1 dw)	Bioconcentration Factor (BCF) (dm3 kg−1)	Plant Quality Standard (PQS) (mg kg−1 dw)	Recommended PQS (mg kg−1 dw)
Red algae	Vertebrata fucoides	24.7 (3.8–83.0)	7717	38.6	50
	Furcellaria lumbricalis	18.8 (5.5–56.8)	5880	29.4
	Coccotylus truncatus	24.2 (7.6–39.6)	7548	37.7
	Ceramium diaphanum	26.5 (4.0–72.7)	8287	41.4
Vascular plants	Stuckenia pectinata	11.0 (3.6–29.2)	3431	17.2
	Zannichellia palustris	18.8 (6.7–87.5)	5869	29.3
	Zostera marina	12.2 (3.0–24.4)	3807	19.0
	Seawater	3.2 µg L−1		recommended 5 µg L−1 WQS

) was determined using the procedure proposed for other metals [42].

Based on the mean concentrations of Cu in the tissues of selected species (C_PL) and the mean concentration of Cu in seawater (C_SW), bioconcentration factors (BCF) were determined according to the formula:

$$\mathrm{B}\mathrm{C}\mathrm{F}={\displaystyle \frac{C_{PL}}{C_{SW}}}$$

(6)

A higher bioaccumulation capacity of copper characterises red algae; BCF values remain in the range of 7500 to 8300 dm³ kg⁻¹ (Table 2). The exception is F. lumbricalis, where the BCF value (5900 dm³ kg⁻¹) is comparable to the value determined for Z. palustris representing vascular plants, of which the other two species have lower BCF values, averaging 3500 dm³ kg⁻¹. These differences result directly from the method of uptake of elements from the environment (the entire thallus in the case of red algae and mainly the root system in the case of vascular plants) and the size of the uptake area (much larger in the case of V. fucoides than in F. lumbricalis [49,50].

The next stage was to determine the PQS based on the BCF values determined for the studied species, considering also maximum concentrations and the new WQS value:

$$\mathrm{P}\mathrm{Q}\mathrm{S}=\mathrm{B}\mathrm{C}\mathrm{F}\times \mathrm{W}\mathrm{Q}\mathrm{S}$$

(7)

Based on the PQS value for species that ranged from 17 to 83 mg kg⁻¹ dw, a universal PQS value for macrobenthic plants equal to 50 mg kg⁻¹ dw was determined, the adoption of which means, taking into account the average concentrations of Cu in the studied species, the achievement of good environmental status in all areas (Table 2).

Fish and shellfish are essential parts of the diets of seafood-dependent countries [51]. Correspondingly, fish and shellfish are well-known indicators of heavy metal pollution [45]. Examining the metal concentration in fish and shellfish meat is important to ensure food safety regulations and consumer protection compliance. Once metal assimilates in biota, it tends to biomagnify in the food chain and becomes difficult to break down into less dangerous compounds [52]. Currently, the permissible concentration of copper in commercially exploited fish is not limited by EFSA (there is only defined an acceptable dose from all sources), unlike metals that have no positive function in the body, such as mercury, cadmium, lead, or arsenic.

The Acceptable Daily Intake (ADI) for copper from all sources in food is set at 0.07 mg/kg bw based on an updated evaluation of scientific evidence. Based on the concentration of Cu in the fish tested in 2022 and assuming consumption of 2.88 kg of fish per year and the average body weight of an adult—70 kg, EDI was calculated (

Table 3. Estimate of Daily Intake rate (EDI) determined for copper concentrations in fish muscle in 2022 with reference to the Acceptable Daily Intake (ADI) value set by The European Food Safety Authority (EFSA) in 2022.

Species	Fishery	Cu [mg/kg]	EDI	Target Consumption [g/day]	ADI	Target Cu [mg/kg]
herring	LWLA	1.79	0.20	2.41	0.07	0.63
	LKOL	2.27	0.25
flounder	Pomeranian Bay	16.12	1.77	0.50
	Gdańsk Bay	3.85	0.42
perch	LZSZ	3.53	0.39	1.78
	LZW	1.98	0.22

). For herring, the mean EDI value was 0.20 mg/kg bw; for flounder, the mean value was calculated to be 1.06 mg/kg bw, while for perch, the respective value was 0.61 mg/kg bw (Table 3). In the case of fish from fisheries in the southern Baltic region, the consumption of fish alone would cause the acceptable dose of copper to be exceeded by the human body with a specific intake and average body weight and was higher than recommended ADI from all sources: 0.07 mg/kg bw (Table 3). Considering the recommended dose for consumption and the statistical consumption of fish, the recommended, safe concentration of copper in fish is a value not exceeding 0.63 mg kg⁻¹ ww, which can be considered a threshold value used to assess the state of the environment. This value has not been achieved in any of the areas, but the values are very close in two areas (the Władysławowo fishery and the Vistula Lagoon). The worst situation is in the case of a flounder from the Pomeranian Bay (Table 3). Therefore, daily intake of copper from fish may have a harmful effect on public health, especially benthic fish. Assuming the consumption of fish to be the only source of Cu in the human body, the safe ADI level would be exceeded by an intake at about higher than 0.50 g per day in the case of flounder muscles to higher than 2.41 g per day by eating herring muscles (Table 3). This means that the consumption of 2–3 servings (1 serving equals ca. 100 g) of fish per month does constitute a threat.

Considering the copper levels in the bottom sediment layers formed during the period of limited anthropogenic pressure (before the period of intensive industrial development, i.e., before 1850), the sediment quality standard (SQS) was proposed at 35 mg kg⁻¹ dw. Because of the total organic carbon (TOC) content in sediments on copper bioavailability, the SQS value was normalised to 5% TOC, resulting in a final value of 40 mg kg⁻¹ dw, which is in high agreement with the copper content in the Earth’s crust (45 ppm). Previous studies based on toxicity analysis recommend using a threshold value for bottom sediments equal to 52 mg kg⁻¹ dw at 5%TOC [53,54]. This means that the value proposed in this study is more restrictive. Taking this value as the boundary, good status, assessed based on the current Cu concentrations in sediments, was achieved only in the eastern Gotland Basin and the Vistula Lagoon, while in other areas, Cu concentrations exceed the proposed SQS value.

4. Conclusions

Studies covering key elements of the marine ecosystem based on current and historical data obtained as a result of long-term analyses of copper in samples and through the use of the dating method made it possible to assess the current situation in terms of Cu concentrations in commercially used fish and benthic plants and in surface bottom sediments, as well as to analyse the temporal variability of copper levels in relation to changes in its inflow to the of the southern Baltic Sea. The complex indicator was developed to assess the state of the environment covering multiple aspects of the entire marine ecosystem (distribution in the water column as a result of the introduction of Cu to the marine environment, bioaccumulation in organisms (fish and plants) and accumulation in bottom sediments. Based on the current Cu concentrations in the studied elements and using the proposed threshold values, which can be used for assessments carried out in accordance with the requirements of MSFD, it should be concluded that good environmental status has been achieved in the case of seawater and plants and has not been achieved in the case of sediments and fish for consumption.

The average concentrations of Cu in herring, flounder and perch in 2022 range from 1.89 mg kg⁻¹ ww to 3.85 mg kg⁻¹ ww. A much higher value was identified in the case of flounder from the Pomeranian Bay (16.12 mg kg⁻¹ ww), which is also associated with an increase in concentrations over time in this case, in contrast to other areas where a gradual decrease in Cu levels in fish is observed. A comparison of the current concentrations to the proposed threshold value (1.86 mg kg⁻¹ ww) indicates that the concentration of Cu at the current level may pose a threat to the potential consumer, which would indicate the need to take further measures to limit the inflow of copper from various sources. The results of the micronucleus test as a method of assessing environmental genotoxicity showed that the limit value (0.39) was not exceeded in 2014 and 2022 for herring from both fisheries and in 2022 for perch from the Vistula Lagoon.

Mean Cu concentrations at the level of 16–32 mg kg⁻¹ dw were characterised by red algae and were higher than those observed in green algae and vascular plants, which is related to the method of uptake of elements from the environment and plant morphology. At the same time, it was shown that the efficiency of bioaccumulation strongly depends on the temperature of the seawater.

The highest concentrations of Cu at the level of 50 mg kg⁻¹ dw in contemporary bottom sediments were recorded in the Gdańsk Basin, which is under the influence of inflowing river waters. Good environmental status was not achieved in the Bornholm Basin, the Gdańsk Basin and the Szczecin Lagoon, where Cu levels exceeded the proposed threshold value (40 mg kg⁻¹ dw).

The study showed that climate change, the main feature of which is an increase in seawater temperature, has a significant impact on the distribution and levels of copper in individual elements of the marine environment. It extends the growing season, influences the bioaccumulation efficiency and is crucial for copper toxicity to marine organisms.