Adverse effects of Dichlofluanid on neotropical marine organisms

doi:10.21203/rs.3.rs-4974125/v1

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Adverse effects of Dichlofluanid on neotropical marine organisms

https://doi.org/10.21203/rs.3.rs-4974125/v1

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This study aimed at evaluating the chronic and acute toxicity of dichlofluanid, an antifouling biocide, on marine invertebrates. Chronic toxicity tests with waterborne dichlofluanid were conducted with embryos of the sea-urchin Echinometra lucunter and ovigerous females of the copepod Nitocra sp. Chronic and acute toxicity tests with sediments spiked with dichlofluanid were performed with Nitocra sp. and the amphipod Tiburonella viscana, respectively, considering sediments with low and high quantities of organic matter. Toxicity was above the environmental concentrations reported in the literature for water. Toxicity to embryos of E. lucunter occurred from 1 µg/L, with effective concentrations to 50% organisms (EC50-40h) estimated as 198.5 (84.6-466.1) µg/L. For the copepod, toxicity started at 100 µg/L, and EC50-7d was 566.4 (304.0-738.4) µg/L. No sediment toxicity was observed to Nitocra sp., while significant effects on the amphipods occurred from 1000 ng/g. Sediments with high organic matter were more toxic. Dichlofluanid may be toxic to marine invertebrates; however, in sediments, its toxicity may be influenced by the amount of OM, since organically enriched sediments exhibited higher toxicity.

Dichlofluanid

toxicity test

water

sediment

antifouling

biocide

Antifouling paints have been employed as a strategy to protect and prevent the establishment of biofouling communities on hard surfaces immersed in seawater [1]. They are used in artificial structures, such as aquaculture tanks, vessels, submarine pipelines, and oil rigs [2]. These paints are used to preserve such structures and maintain navigability in the case of vessels [3].

Anti-fouling compounds include various organic and inorganic substances and many of them are potentially harmful to aquatic ecosystems, as they are planned and designed to be toxic to the aquatic biota and normally leach from the painted surfaces to the water column [4]. Regions with an intensive presence or flow of vessels, such as ports, shipyards, hull zones, and marinas, are often susceptible to contamination by this type of chemical substance [5]. In addition, physical, chemical, and biological processes contribute to the transfer of these substances from water to sediments and biota [3–5], as they are often associated with organic matter or mud particles. On the other hand, other processes may cause the resuspension of particles from the bottom and the consequent remobilization of chemicals back to the water column, such as bioturbation, dredging, and storms [6, 7]. Antifouling paints leached from vessels to water column or sediments can be subject to a range of processes, such as biodegradation, bioturbation, sorption onto particulate matter, photolysis, and volatilization, all of which can interfere with the fate and effects of anti-fouling biocides after they are released from painted surfaces [7, 8].

The first generation of antifouling paints began to be used in the mid-20th century and contained oxides of copper and zinc as their active principles [9, 10]. These paints were considered inefficient and their active principles were substituted by a second generation of biocides [1, 9, 11] which were represented mainly by organotin (OT) compounds, especially tributyltin (TBT) and/or triphenyltin (TPhT). These new paints rapidly became widely used in vessels worldwide during the 1980s. OT-based coatings were extremely efficient but highly toxic, being capable of causing adverse effects on non-target species. This caused them to be banished from the composition of antifouling paints by the International Maritime Organization (IMO) [3]. In September 2008, the traffic of vessels traffic covered with OT paints was definitively banned [1, 12], at least for those countries that complied with international treaties. As a consequence, the third generation of antifouling paints was released in the market [3, 13], consisting of paints containing a range of chemicals, either alone or in combination, in their formulation. Up to four antifouling compounds can be used for the same commercial product [1, 13]. These compounds belong to different chemical classes and include non-metallic compounds such as Chlorothalonil, Dichlofluanid, Diuron, Irgarol 1051, Sea-Nine 211, Pyridine (TPBP and TCMS), Thiram, and TCMTB, as well as metallic or organometallic compounds such as Maneb, Zineb, Ziram, zinc pyrithione, copper pyrithione, copper thiocyanate, copper naphthenate, and copper oxide [1, 14]. Irgarol and Diuron are further phased out of antifouling paints because of their persistence and toxicity [15–17].

Dichlofluanid (N-[dichloro(fluoro)methyl]sulfanyl-N′-(dimethylsulfamoyl) aniline) is a third-generation antifouling biocide that is primarily used as a fungicide to control diseases on fruits and ornamental plants [1, 18] and to function as a wood preservative [19], and it has recently been used as an antifouling agent. Dichlofluanid has been detected in surface sediment samples collected in the field [20, 21]. This compound is an organochlorine with low solubility (0.006 mg/L) and a Kow of 3.7, and these properties provide dichlofluanid a high affinity for sediment particles and particulate matter [1, 4, 5]. It has a short half-life in aqueous mean (i.e., < 24 h) and water-sediment systems (between 1.2 and 3 h) [20], and after being released into the water, Dichlofluanid tends to rapidly degrade to N,N-dimethylaminosulfanilide (DMSA), which is not easily biodegradable in water/sediment systems [22] but abiotically may generate other byproducts, such as dichlorofluormethane and N,N-dimethylsulfamide (N,N-DMS) [14, 23–25]. Particularly, N,N-DMS is a persistent and mobile organic compound (PMOC) and may remain for longer periods in sediments [22, 26–27]. The degradation of dichlofluanid is influenced by the quantity of organic matter, pH, incidence of light, and microbial activity [28]. Dichlofluanid is an extremely toxic compound, whereas DMSA is less toxic in aquatic environments [20]. According to the Danish Environmental Protection Agency [29], there is very little data on the toxicity of N,N-DMS, and other transformation products from dichlofluanid.

Studies conducted in different regions, such as Europe and Asia, showed that environmental concentrations of dichlofluanid in sediments could range from < 1.6 ng/g to 800 ng/g [30–34]. In water, the concentrations found ranged from < 4 ng/L to 600 ng/L in Central America, North America, Asia, and Europe [20, 30, 31, 35–39]. In Brazil, sediment concentrations of 16 ng/g were found in Santos Bay, and in Espirito Santo Bay, concentrations ranging from < 0.7 to 6.5 ng/g [40–42]. However, there is still a need to determine the toxic levels of dichlofluanid in marine organisms, as this information is critical for the estimation of risk and hazard assessments and the establishment of regulations for this compound.

To evaluate the ecotoxicological risks associated with dichlofluanid, tests with aqueous solutions and spiked sediments are necessary because both water and sediment are critical environmental compartments. The water column is the first medium to receive antifouling compounds leached from painted surfaces, while the sediments can accumulate contaminants over time [43]. Moreover, sediments may have variable quantities of organic matter (OM), and this factor is worth studying because OM is capable of binding, adsorbing, and regulating the bioavailability of organic compounds due to its affinity for hydrophobic compounds [44, 45]. The interactions between organic compounds and OM include processes such as ion exchange, covalent binding, hydrogen bonding, hydrophobic adsorption and partitioning, and charge transfer [46–49], which can greatly influence the toxicity and bioavailability of organic compounds in marine organisms [50].

Toxicity tests have been routinely employed to determine and/or assess the effects of xenobiotics in the marine environment [1, 51]. However, only a few studies have assessed the toxicity of dichlofluanid in marine organisms [1]. This study aimed to assess the toxicity of dichlofluanid to marine organisms at different concentrations, including realistic environmental concentrations. In particular, we aimed to assess the toxicity of dichlofluanid in aqueous solutions and sediments and to evaluate the influence of OM on the toxicity of dichlofluanid.

2.1. Toxicity tests with aqueous solutions

The toxicity tests of dichlofluanid in water were performed at the following concentrations: 0.1, 1, 10, 100, and 1000 µg/L (nominal concentrations). Dichlofluanid is a biocide that generally has an extremely short half-life and rapid degradation; therefore, nominal concentrations were used because this substance rapidly transforms into other compounds whose toxicities are variable or unknown [1].

The working solutions were prepared at the time of the experiments by dilutions of the stock solutions in filtered seawater with the salinity adjusted for each specific toxicity test. A stock solution (1 g/L dichlofluanid) was prepared in acetone. The negative controls consisted of clean seawater (collected off the coast in a region free of pollution) and were also prepared in each experiment. Besides, "analytical blanks" solution containing only acetone solvent diluted in seawater were prepared, at concentrations ranging from 0.05–0.1% of the total exposure concentration. At this concentration range, acetone is reported to be not toxic to the organisms used in this research [52, 53].

In all the tests performed, physicochemical parameters were measured to ensure that the test conditions remained within the acceptable ranges for the respective test species, and thus, any variations in these variables would not interfere with the results. Salinity, pH, temperature, and levels of dissolved oxygen first were measured using appropriate electrodes, whereas the ammonia concentrations were quantified using a colorimetric method [54]. The results obtained in all tests were compared with the respective controls and analytical blanks.

2.1.1. Embryonic development test – Echinometra lucunter

The embryonic development tests using eggs and larvae of the sea urchin Echinometra lucunter were conducted according to the protocol proposed by ABNT [55], which is an adaptation of international protocols [56, 57].

Adult sea-urchin individuals were collected by snorkeling at Palmas Island, Guarujá, São Paulo (24° 00'S − 46° 27'W) and immediately transferred to plastic boxes with seawater. They were transported to the laboratory and then placed in tanks containing filtered seawater, where they were kept at constant conditions (25 ± 2°C; photoperiod 12h:12h light-dark). The spawning was induced by osmotic induction, which consisted of injecting up to 3 mL of 0.5 M KCl into the organism's coelomic cavity. Ovules of each female were separated and examined for their morphology and viability, and sperm activation was made through dilution in seawater (0.5 mL sperm into 24.5 mL clean seawater). Fertilization was made by adding approximately 2 mL of the sperm solution into the ovules solution. The eggs solution was then left agitating for 2 hours, and then, the fertilization success was checked: three sub-samples of the eggs solution were taken, counted, and examined for the presence of the fertilization membrane. The fertilization success was confirmed when at least 80% of eggs exhibited the fertilization membrane. Then, approximately 500 eggs were introduced in each test chamber, which consisted of 15 mL glass tubes containing the dichlofluanid solutions and the respective controls and analytical blanks. Four replicates were prepared for each treatment. The test was kept in an incubator for approximately 40–44 hours (time required for the formation of the equinopluteus larvae in the follow-up controls) at 25 ± 2° C, under a 16h:8h light/dark photoperiod. At the end of the test, the content of each replicate was then fixed by the addition of tamponed formaldehyde (10%). Afterwards, 100 embryos were counted from each replicate, by the use of an optical microscope (100X magnification) and the percentages of normal embryos were calculated. Normal embryos were defined based on typical larval development, which included the branch symmetry, and shape and size of the skeleton [58].

2.1.2. Fecundity test – Nitocra sp.

The chronic toxicity assay using the benthic harpacticoid copepod Nitocra sp. followed the standard protocol developed by [59, 60], which was an adaptation of the protocol adopted by the US Army Corp of Engineers [61]. Nitocra sp. has a relatively short life cycle and is highly sensitive to contaminants [1]. Besides, this organism is estuarine, benthic, cosmopolitan, and easy to cultivate in the laboratory [59, 60]. The organisms used in this investigation have been reared in the laboratory since 1999.

Four replicates were prepared in 30-mL glass beakers, each containing 20 mL dichlofluanid solution, or the respective controls and analytical blanks. Ten healthy ovigerous females were introduced into each test chamber. The whole test system was maintained at 25 ± 2º C for 7 days with a photoperiod of 16h:8h (light:dark), and the salinity was kept at 17 ± 2. At the end of the tests, the content of each test chamber was fixed by adding tamponed formaldehyde (10%) and Rose-Bengal dye (0.1%). Then, the numbers of adult females and offspring (nauplii and copepodits) in each replicate were counted using a stereomicroscope.

2.2. Toxicity tests with spiked sediments

Sediments of two types were used in this study; they differed by the respective concentrations of organic matter (OM): low and high quantities of OM. These sediments were collected at two distinct sites in the Cananéia region [1], within the Lagamar Protected Area, on the south coast of the state of São Paulo. The sediment containing low quantities of OM (low OM) was collected at Arrozal (25°02.415’ S and 47°55.540’ W), while the sediment with higher OM amounts (high OM) was collected further south, at Ariri (25°13.215’ S and 48°02.492’ W). One of the tests performed with the low OM sediment (Tiburonella viscana test) was repeated because of operational problems in the laboratory. Thus, the results of sedimentological properties are further presented for three sediments: one sample collected in Ariri (High OM) and two collected at Arrozal, at different moments (Low OM and Low OM_2).

Sediments were spiked with dichlofluanid, considering the quantities (in %) of OM. Similar to the tests with aqueous samples, the negative controls and analytical blanks were prepared with the same sediments. The spiking procedure is further described in this article. In all toxicity tests performed, physicochemical parameters in the overlying water within the test chambers were monitored, using the methods described in the previous section (i.e., tests with aqueous solutions).

2.2.1. Sedimentological analyses

The sediment grain size distribution was analyzed according to the method described by Mudroch and MacKnight [62]. In both the measurements and experiments, the calcium carbonate was not removed because it is an integral part of the sediment; its removal could have potential geochemical, ecological, and ecotoxicological implications. Three 30 g aliquots of sediment were separated and oven-dried at 60 ºC for 48 h. The sediments were then washed in a 0.063 mm mesh sieve to separate fine particles (mud and clay). The sandy fraction retained on the sieve was dried in an oven at 60° C for more than 48 h. This material was weighed and sifted for 15 minutes in an RO-TAP shaker, passing through different mesh sizes (ɸ scale) to separate the different classes of sand. The results were classified using the Wentworth Scale [54], and the average of three samples was calculated.

The quantities of calcium carbonate (CaCO3) in each sample were quantified using the acid digestion method [64]. Three fractions (5 g) of each sediment were separated in glass beakers, and 10 mL of 5N HCl were added to each beaker, for 24 h, to eliminate calcium carbonates. The remaining sediments were washed with distilled water, dried in an oven (at 60°C), and weighed again. The weight difference between initial and final measurements allows the calculation of the percentage of calcium carbonate present in the sample. This procedure was repeated three times for each sample, and the average was calculated.

The quantities of organic matter (OM) in the sediments were estimated using the loss-by-ignition method [65]. It consists of separating three samples (5 g) of dry sediment from each sample incinerating them in a muffle (500°C) for 4 h and then weighing them again. The mass loss after incineration allows the calculation of OM percentage. The average of three samples was calculated, similar to CaCO₃.

2.2.2. Sediment spiking procedures

Five concentrations of dichlofluanid were used in the tests with spiked sediments: 1, 10, 100, 1000, and 10000 µg/g. The sediment spiking procedure consisted of that described by the ASTM [66] and Fuchsman and Barber [67]. The dichlofluanid at the appropriate concentrations were added to the sediment, taking into consideration the dry weight of the sediment. The wet sediment was split into aliquots and different quantities of dichlofluanid were added, to produce the test concentrations. The control and analytical blank sediments were submitted to the same procedure. The sediments were shaken for 15 min using a Jar test, which consisted of multidirectional rotations of the sediment samples in a hermetically sealed glass container. After this mixing procedure, the spiked sediments were stored in the dark, at 4°C, for 72 h, to allow the establishment of an equilibrium state between the interstitial water and the solid fraction of the sediment. The sediments were then distributed in the test chambers for toxicity tests. Details of the procedures used in sediment spiking are available in Campos et al. [68], as well as the chemical confirmation of dichlofluanid concentrations in the sediments analyzed in this study. These authors quantified the dichlofluanid using Gas Chromatography coupled with Mass Spectrometry (GC-MS) techniques.

2.2.3. Acute sediment toxicity test – Tiburonella viscana

The acute sediment toxicity test with T. viscana was conducted following the protocol described by Melo and Abessa [69] and ABNT [70]. The amphipods used in this test were collected at Engenho d'Água Beach, Ilhabela, São Paulo, Brazil (23°48' S – 45°22' W). Three replicates of each sediment sample were prepared, in glass chambers containing 150 g of sediment and 600 mL of clean seawater (salinity = 35) [1]. Ten healthy, non-ovigerous adult amphipods were introduced into each replicate. The test was conducted for 10 days, under constant lightning, aeration, and temperature (25 ± 2º C). After 10 days, the contents of each replicate were sieved and the surviving organisms were separated and counted. Missing organisms were considered dead.

2.2.4. Chronic sediment toxicity test – Nitocra sp.

Before the chronic toxicity tests using copepods, we conducted preliminary experiments to verify whether the natural sediments (low and high OM) could cause any toxicity to the organisms [1]. The results of these experiments are shown in the Supplementary Material (Tables SM-1 to SM-5), and they did not show any evidence of toxicity in both sediment types. In the chronic sediment toxicity tests, ovigerous females were separated from the cultures and exposed to sediments containing different concentrations of dichlofluanid for 7 days, and then the fecundity rates were evaluated. The tests were conducted following the protocols proposed by Lotufo and Abessa [59] and ABNT [60]. The procedure was relatively similar to that described for aqueous samples, except for the presence of sediments in the test chambers, as approximately 4 g of spiked sediments (or the control or analytical blanks) and 16 mL of water at a salinity of 17. Four replicates were set up for each treatment.

2.3.Statistical analyses

The results obtained in the water toxicity tests were initially verified for normality and homoscedasticity using Shapiro Wilk and Bartlett tests, respectively. If they passed, a one-way ANOVA was performed, followed by Dunnet's test (or its equivalent nonparametric test) to identify significant differences between the treatments (dichlofluanid spiked sediments), analytical blanks, and the controls. The lowest observed effect concentration (LOEC) and no observed effect concentration (NOEC) were calculated from this data. Effective Concentrations (ECs) were also for 50% and 10% of the exposed organisms (EC50 and EC10, respectively) based on the linear interpolation method, using the ICPIN software. The results of the tests with spiked sediments were also checked for normality and homoscedasticity, but they were analyzed by a two factors ANOVA, to identify significant differences between the treatments, analytical blanks, and controls, as well as the two types of sediments (low and high MO). These analyses were conducted using the GraphPad Prism software.

3.1. Toxicity of aqueous solutions

The percentages of E. lucunter abnormal larval development after exposure to a range of dichlofluanid concentrations are displayed in Fig. 1. The physicochemical parameters of the test solutions remained within the accepted limits for the test, as stated by [55] (Table SM-1). A significant increase in the percentage of abnormal larvae was observed from 1 µg/L; thus this concentration was the LOEC, and consequently, the NOEC was 0.1 µg/L. The EC50-40h was calculated as 198.5 (84.56–466.1) µg/L whereas the EC10-40h was 0.36 (0.01–12.02) µg/L. The embryos exposed to 1000 µg/L showed a 100% abnormality rate.

The results of the toxicity tests using harpacticoid copepods and dichlofluanid are shown in Fig. 2. The physicochemical parameters of these tests are shown in the table SM-2. In the first experiment (Fig. 2A), at concentrations similar to the environmental levels, no toxicity was observed, as there was no difference between the offspring development of the exposed organisms and the acetone control (ANOVA, p = 0.32). In the other experiments (Fig. 2B, 2C, and 2D), performed with higher concentrations, Nitocra sp. exhibited a significantly lower fecundity at 1000 µg/L of dichlofluanid, when compared to the 0.05% acetone control (Fig. 2B and 2C). In the fourth test, significant effects occurred from 100 µg/L (Fig. 2D). In the third test, a significantly smaller number of F1 (copepodits + nauplii) was observed in the water control group. This was likely caused by the fact that the mean number of offspring in the acetone control 0.05% was slightly above the rate normally observed in tests with aqueous solutions in our laboratory. The LOECs varied between 100 and 1000 µg/L while the NOECs ranged from 10 to 100 µg/L, depending on the test. Based on the values obtained in tests 2, 3, and 4 (Figs. 2B, 2C, and 2D) the mean EC10-7d was 0.06 (0.04–70.9) µg/L, while the mean EC50-7d was 566.4 (304.0-738.4) µg/L.

3.2. Toxicity of spiked sediments

3.2.1. Sedimentological analyses

The grain size distribution and the percentages of CaCO₃, OM, and dry weight of the sediments tested are shown in Table 1. Both sediments were predominantly composed of mud and fine and very fine sands. Thus, any possible modifications in toxicity would not be caused by sediment properties. The sediment from Ariri presented a higher quantity of OM (10.4%) than the sediment of Arrozal, which had approximately 4% OM. Thus, they were respectively named as high OM, low OM, and low OM_2. Finally, regarding the quantities of CaCO₃ in the sediments, no significant differences were detected; thus this factor possibly did not influence the results of the toxicity tests.

Table 1

Percentages of grain size distribution, organic matter (OM), Calcium Carbonate (CaCO₃), and wet weight of sediment in the different sediments collected in Cananéia, (SP, Brazil) and used in toxicity tests with dichlofluanid.
	Gravel	Sand					Silt & Clay	OM	CaCO₃	Wet Weight
	Gravel	Very Coarse	Coarse	Medium	Fine	Very Fine	Silt & Clay	OM	CaCO₃	Wet Weight
High OM	0.37	0.33	2.11	5.21	18.22	55.63	17.10	10.38	7.38	57.18
Low OM	0.00	0.07	0.62	1.42	8.66	74.53	13.81	4.00	5.88	49.52
Low OM_2	0	0.03	0.04	2.04	48.78	43.29	5.81	4.00	1.24	20.43

OM: Organic Matter; CaCO₃: Calcium carbonate.

3.2.2. Chemical confirmation

As mentioned, the chemical confirmation of dichlofluanid was evaluated by Campos et al. [68], as shown in Table 2. Dichlofluanid was only detected at 1000 n/g, with degradation rates of 87% and 99% at 6 and 24 h, respectively. Dichlofluanid is a biocide that generally has an extremely short half-life [20], with rapid degradation. However, because dichlofluanid has been reported at concentrations above its toxic threshold, and the degradation of dichlofluanid may generate sub-products that can have different toxicities, this study is important.

Table 2

Concentration of dichlofluanid in sediments at 0, 6, and 24 hours after spiking under equilibrium condition, as obtained by Campos et al [68].
	Nominal concentratiom (ng/g)	Time
	Nominal concentratiom (ng/g)	0h	6h	24h
Dichlofluanid (LOD = 0.5; LOQ = 1)	Blank	<LOD	<LOD	<LOQ
	10	<LOD	<LOQ	<LOQ
	100	<LOQ	<LOQ	<LOQ
	1000	1189.5 ± 460.4	157.7 ± 36	16.9 ±7.4

LOD = Limit of Detection; LOQ = Limit of Quantification.

3.2.3. Sediment Toxicity

Initially, a chronic toxicity test was carried out with the copepod Nitocra sp. to assess whether the collected sediment would exhibit a decrease in fecundity before contamination. The collected sediments were compared with sediments frequently used as reference in the laboratory. The results of the test and the physicochemical parameters are shown in the supplementary material (Figure SM-1 and Table SM-3). The results of the spiked sediment tests with Nitocra sp. and Tiburonella viscana are presented in Figs. 3 and 4, and the physicochemical parameters of the overlying waters in the test chambers are shown in tables SM-4 and SM-5, respectively.

In the test with Nitocra sp., there was no reduction in the fecundity of the organisms compared to the acetone control. When high and low OM sediments were compared, no significant differences were observed at the same concentrations. The low OM sediments did not cause significant effects in T. viscana at any of the tested concentrations. However, the high OM sediments caused toxicity at all the concentrations tested, when compared to the acetone control.

Significant differences between the acetone control and all the test concentrations could not be properly detected by parametric or not-parametric statistical tests because the standard deviation of this treatment was equal to zero; in such cases, any difference in the mortality rates would be considered significantly different from the control by the statistical tests. Mortality rates increased up to 40% at 1000 ng/g and 30% at 10000 ng/g, while lower concentrations produced lower mortality rates. When the high and low OM sediments were compared, the concentration of 1000 ng/g presents the greatest difference between the mortalities; thus, it can be inferred that organic matter can influence the toxicity of dichlofluanid to T. viscana, and this influence can be perceived from this concentration.

The experiments showed that dichlofluanid affected the embryonic development of E. lucunter and the fecundity of Nitocra sp, corroborating other studies that have shown that dichlofluanid may cause toxicity to different aquatic species. Bellas [71] evaluated the effects of dichlofluanid on the embryo-larval development of the sea-urchin Paracentrotus lividus, the bivalve Mytilus edulis, and the ascidian Ciona intestinalis, and showed that concentrations between 17 µg/L and 1035 µg/L could increase the rates of abnormal embryo-larval development. Tests with embryos of the sea urchin Glyptocidaris crenularis showed an increase in abnormal embryonic development at the highest concentrations (500 µg/L and 1333 µg/L, respectively) and obtained an EC_50-50h of 0.41 µM (92 µg/L) [53]. The respective EC₅₀ and EC₁₀ were 1000 and 11.7 µg/L for the copepod Acartia tonsa [72], while for Crassotrea gigas, the same authors observed an EC₅₀ of 131 µg/L and an EC₁₀ of 98.7 µg/L. Previous studies conducted with Daphnia magna showed that the lethal effects on 50% of organisms (LC_50-48h) were 370 µg/L [73] and 1800 µg/L [74]. The above-mentioned studies with different aquatic species showed effects of the same order of magnitude as we found the effects on marine invertebrates; thus, the toxicity ranges obtained in the present study are consistent in comparison with the results reported in the literature

The toxicity of dichlofluanid can also be compared to that caused by other antifouling biocides. Compared to the current study, dichlofluanid appears to be less toxic than bis-tri-n-butyltin(IV)oxide (TBTO), which had an LC_50-96h of 13 µg/L to the copepod Nitocra spinipes [75], and more toxic than Irgarol and Diuron, which presented an LC_50-96h of 4500 µg/L to N. spinipes [75]. Perina et al. [58] studied the toxicity of some biocides to embryos of the sea urchin Lytechinus variegatus and reported that for TBT and TPT the NOEC and LOEC were 0.25 µg/L and 0.5 µg/L while for Diuron and Irgarol the NOEC and LOEC 1000 µg/L and 2000 µg/L respectively. Based on these data and our results, we can infer that TBT and TPT are likely more toxic than dichlofluanid, while Diuron and Irgarol are less toxic.

In general, considering the exposures performed in aqueous solution, the concentration of 1000 µg/L caused more consistent toxic effects, such as embryonic abnormalities and reduction of copepod reproduction. This concentration was above the highest environmental concentrations found worldwide [20, 30–36, 76]. Moreover, dichlofluanid rapidly degrades when aqueous solution. At pH > 7, this compound is rapidly hydrolyzed to dimethylaminosulfanylide (DMSA) [77], followed by N,N-DMS, and the toxicity of these by-products can differ from that of dichlofluanid [20]. In our study, dichlofluanid dissolved in clean seawater did not cause significant adverse effects in the organisms studied at realistic environmental concentrations, at least in the short-term exposures considered herein. However, the N,N-DMS, a persistent compound with unknown toxicity, is not susceptible to degradation and may have a half-life of > 1000 days in the water-sediment system [22, 78]. Thus, the use of this biocide as an antifouling compound is concerning because of ecotoxicological implications [1, 22].

In sediments, organic substances tend to partition into organic matter and establish a range of chemical interactions that directly influence their bioavailability and potential toxicity [45, 79]. In this context, biodegradation and volatilization may affect the persistence of organic contaminants in sediments [79, 80]. In the current study, dichlofluanid may have been absorbed by benthic organisms via the food route, to the point of inducing toxic effects. Organisms such as amphipods and harpacticoid copepods ingest particles of organic matter and contaminants when feeding on sediment, being susceptible to contaminated food [1].

The measured environmental concentrations (MEC) of dichlofluanid in sediments from Europe and Asia ranged from < 1.6 to 688.2 ng/g [30–37]. Based on these MECs, we inferred that toxicity could be expected in the environment since our data showed toxic effects at low concentrations. The adoption of adjustment factors aims to allow estimation of the effects generated after long-term exposure and calculation of environmental hazards and risks, so that even low environmental concentrations may be concerning. In this study, adverse effects could be expected at concentrations equal to or above 1000 ng/g, which would be considered as the LOEC. Considering that a concentration of 1 ng/g would be the lowest available NOEC and based on the Technical Guidance Document on Risk Assessment [81], it is possible to estimate the Predicted No Effect Concentration (PNEC) using a safety factor of 10. In the present study, the PNEC value was 0.1. Therefore, adverse effects on the ecosystem are predicted because the environmental concentration of a contaminant is greater than that of the PNEC. Abreu et al. [40] showed that for Brazil, environmental concentrations of 16 ng/g were observed. Thus, environmental risks due to dichlofluanid could be expected in sediments containing high quantities of OM. In a previous study using risk quotients [71], the authors concluded that dichlofluanid could present a risk to some aquatic species.

The impact assessment of pollutants needs to consider that, in most cases, contaminants occur in complex mixtures in the environment. Xu et al. [53] observed that, when biocides of third-generation antifouling paints and heavy metals were combined in binary mixtures, the effects were synergistic. The authors also reported that dichlofluanid caused worse toxic effects on sea urchin embryos when combined with metals. Another study [82] evaluated the interactions between the biocides of the third generation and revealed that such interactions could cause a synergistic toxic response, suggesting that the toxicity of individual compounds alone was not sufficient to determine the environmental impacts of toxicants [82]. In any case, the toxicity of single compounds is necessary for the establishment of due regulations and comprehension of their potential to cause adverse effects to the biota, together with other approaches. In the present study, dichlofluanid adversely affected the tested organisms, and under the light of data from the literature, it seems to be less toxic than some compounds formerly used as biocides, such as the TBT. However, at the same time, dichlofluanid seems to be more toxic than other common antifouling agents such as Irgarol and Diuron [15, 17, 58, 71, 75].

This study shows that, at environmental concentrations, dichlofluanid in an aqueous solution likely does not cause significant adverse effects on the embryonic development of E. lucunter and the fecundity of Nitocra sp. However, at 1000 µg/L, dichlofluanid can affect the larval development of E. lucunter and also reduces fecundity and survival of Nitocra sp. On the other hand, in sediments organically enriched (i.e., high OM), dichlofluanid caused acute toxicity from 1000 ng/g, whereas sediments with low amounts of OM were not toxic. Thus, results obtained in this investigation suggest that OM may reduce the degradation of dichlofluanid (or its by-products), thereby increasing its toxicity and environmental risks when present in organically rich sediments.

Funding: This study was supported by Financiadora de Estudos e Projetos - FINEP (Process 1111/13–01.14.0141.00), and Conselho Nacional de Desenvolvimento Científico e Tecnológico – CNPq (Grant #456372/2013–0). ACFC (Grants #2016/02029-4 and #2021/06167-0), B.G.C. (Grant #2017/10211–0), and GFEP (Grant #2018/23279-4) were funded by the Fundação de Amparo à Pesquisa do Estado de São Paulo (FAPESP). D.M.S.A. (PQ #308533/2018–6 and #313420/2023-8) and IBG (PQ #304398/2021-7) are research fellows of CNPq. The APC was waived by the journal publisher.

Conflicts of Interest: The authors declare no conflicts of interest.

Data Availability Statement: All data generated or analyzed during this study are included in this published article and its supplementary information files..

Author Contributions: Conceptualization, A.C.F.C. and D.M.S.A; methodology, A.C.F.C., and R.C.K.; validation, A.C.F.C. and D.M.S.A.; formal analysis, A.C.F.C. and R.C.K.; investigation, A.C.F.C.; resources, A.C.F.C. and D.M.S.A.; data curation, A.C.F.C.; writing—original draft preparation, A.C.F.C.; writing—review and editing, D.M.S.A., R.C.K., G.F.E.P., B.G.C., and Í.B.C.; supervision, D.M.S.A. and Í.B.C.; project administration, A.C.F.C. and D.M.S.A.; funding acquisition, D.M.S.A. All authors have read and agreed to the published version of the manuscript.

Institutional Review Board Statement: Ethical review and approval were waived for this study because it involves only experiments with invertebrates, which do not require authorizations or analyses by ethics committees.

Informed Consent Statement: Not applicable.

Acknowledgments: We thank the NEPEA laboratory staff and the UNESP-CLP employees for their support. We are also grateful to Prof. Gilberto Fillmann (FURG) and Dr. Fernando Perina (University of Aveiro) for their assistance, and “Rede Nacional de Estudos de Anti-incrustantes” (RNEA) for the support.

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Adverse effects of Dichlofluanid on neotropical marine organisms

Status:

Version 1

Abstract

Figures

1. Introduction

2. Materials and Methods

2.1. Toxicity tests with aqueous solutions

2.1.1. Embryonic development test – Echinometra lucunter

2.1.2. Fecundity test – Nitocra sp.

2.2. Toxicity tests with spiked sediments

2.2.1. Sedimentological analyses

2.2.2. Sediment spiking procedures

2.2.3. Acute sediment toxicity test – Tiburonella viscana

2.2.4. Chronic sediment toxicity test – Nitocra sp.

2.3.Statistical analyses

3. Results

3.1. Toxicity of aqueous solutions

3.2. Toxicity of spiked sediments

3.2.1. Sedimentological analyses

3.2.2. Chemical confirmation

3.2.3. Sediment Toxicity

4. Discussion

5. Conclusions

Declarations

References

Additional Declarations

Supplementary Files

Status:

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