Toxicity of Antifouling Biocides and Wastes from Ships’ Surfaces during High-Pressure Water-Blasting Cleaning Activities in the Nauplii and Eggs of the Estuarine Copepod Paracalanus parvus sl

Abstract

Copepods, the dominant member of zooplankton and major grazers of phytoplankton in the pelagic ecosystem, are at risk from exposure to antifouling biocides. To evaluate the developmental toxicity of antifouling biocides (Diuron, Irgarol 1051, Sea-nine 211) and wastewater (from high-pressure water blasting (WHPB) and its MeOH extract (WHPB-MeOH)) in the copepod Paracalanus parvus sl, we investigated the chemical concentration, egg-hatching rate, and nauplius mortality. WHPB samples were obtained through hull-cleaning activities involving WHPB in a dry dock. Among the biocides, Sea-nine 211 had the strongest effects on hatching rates and nauplius mortality, which was followed by Diuron and Irgarol 1051. In the WHPB and WHPB-MeOH samples, there was no significant difference between the experimental groups in terms of the egg-hatching rate; however, WHPB was found to be more toxic in terms of nauplius mortality, suggesting that metals in WHPB may also adversely affect nauplius survival in P. parvus sl. A comparison of the LC₅₀ results of Sea-nine 211 and WHPB revealed that WHPB had a negative effect on nauplius mortality even at a 100-fold lower concentration. Therefore, if chemical contaminants generated during in-water cleaning activity are discharged continuously into the ports without being properly collected through a post-treatment system, they are expected to negatively impact the population of copepods near the port. Although verification is needed through additional experiments, our results could be used for a baseline study concerning the toxicity of antifouling biocides on marine copepod species.

1. Introduction

Antifouling paints are applied to the hulls of ships to inhibit the unwanted attachment and growth of fouling organisms such as bacteria, diatoms, barnacles, tubeworms, mussels, tunicates, and seaweeds [1]. The accumulation of these organisms on a ship’s hulls not only leads to increased friction, resulting in higher fuel consumption and cleaning costs, but also induces ecological risks by increasing greenhouse gas emission and introducing invasive species into the native environment [2,3]. Because of the lack of adequate biofouling management regulations, antifouling paints—composed of various substances including pigments, solvents, metals, and organic and organometallic biocides—are released into the coastal and harbor waters during in-water cleaning activities to manage ships’ hulls for repair, cleaning, and painting [4,5,6,7]. Consequently, high concentrations of biocides, booster biocides, and metallic elements have been detected in harbors and coastal areas [8,9,10]. Several investigations on the impacts and risks of antifouling paints on the marine environment have revealed that antifouling paint particles have a toxic effect on the survival and growth inhibition of phytoplankton species [11,12,13]; retard egg production and increase mortality (from larvae to adult) in zooplankton species [13,14,15,16]; and induce developmental defects in sea urchins (Paracentrotus lividus), mussels (Mytilus edulis), and oysters (Crassostrea virginica) [17,18]. However, most studies have focused on the effects of a single biocide and/or metallic compounds on an individual organism; therefore, available data for understanding the toxic effects of more complex antifouling system mixtures are limited [7].

Marine copepods are a diverse and abundant group of crustacean zooplankton, which are an important part of the diet of larger invertebrates, fish, and mammals [19,20]. To obtain information on the ecological effects of chemical components, many research groups are currently investigating copepods as they have short life cycles and fulfill several desirable criteria that are ecologically important. However, toxicity in copepods differs according to gender [21,22,23]. Medina et al. [22] found that Acaritia tonsa adult males are twice as sensitive as females to cypermethrin exposure. Furthermore, early life stages of marine organisms are more susceptible to toxicity than adult life stages [24]. Accordingly, to investigate the developmental toxicity of antifouling systems on the copepod species, we selected Paracalanus parvus sl, which is a dominant copepod species in the neritic waters of South Korea and has a short developmental time (1–2 days) from egg to nauplius.

In this study, we aimed to investigate the toxic effects of wastewater—generated from ships’ surfaces during hull cleaning using high-pressure water jets (WHPB)—on larval development in the early life stage. MeOH extracts (WHPB-MeOH) were used for the selective toxicity of organic compound booster biocide by excluding the toxicity of metallic elements in WHPB. In addition, we investigated the toxic effects of three booster biocides (Irgarol 1051, Diuron, and Sea-nine 211) for a comparative analysis with the toxicity results of WHPB. This study is very meaningful because it is the first attempt at understanding the toxicity of WHPB on the marine copepod P. parvus sl.

2. Materials and Methods

2.1. Test Compounds

The three organic booster biocides used for toxicity tests were Irgarol 1051 (2-(tert-Butylamino)-4-(cyclopropylamino)-6-(methylthio)-s-triazine; CAS No.: 28159-98-0; Sigma-Aldrich, St. Louis, MO, USA), Diuron (3-(3,4-Dichlorophenyl)-1,1-dimethylurea; CAS No.: 330-54-1; Sigma-Aldrich, USA), and Sea-nine 211 (4,5-dichloro-2-octyl-isothiazolone; Sigma-Aldrich, St. Louis, MO, USA). All biocides were dissolved in 99.9% pure dimethyl sulfoxide (DMSO: ACS regent, Sigma-Aldrich, St. Louis, MO, USA) to prepare stock solutions. Stock solutions of Irgarol 1051 (1 g L⁻¹), Diuron (1 g L⁻¹), and Sea-nine 211 (1 g L⁻¹) were diluted with filtered artificial seawater to obtain working solutions at designated concentrations for toxicity tests (

Table 1. Experimental conditions of wastewater (WHPB and WHPB-MeOH) from high-pressure water blasting and three booster biocides (Diuron, Irgarol 1051, and Sea-nine 211).

Species	WHPB	WHPB-MeOH	Diuron	Irgarol 1051	Sea-Nine 211
	(Dilution Gradient)		mg L−1
P. parvus sl	50	1	15	15	1
	100	2	10	10	0.5
	1000	22	7.5	7.5	0.1
	5000	110	5	5	0.01
	7500	155	2	2.5	0.005
	8000	220	1	1	0.001
	10,000	1099	0.1	0.1	0.0001

). The experimental concentration range of each booster biocide was set through a pre-test that was conducted under high, middle, and low concentrations.

WHPB samples (approximately 1.8 L) were collected from the surface of the ship R/V EARDO during high-pressure cleaning of the ship’s hull in a dry dock. The WHPB samples were then stored at 4 °C in a cold box and transported to the laboratory. WHPB-MeOH samples were provided by a co-researcher participating in the research project [7]. The analysis methods for metallic elements and biocides in WHPB samples and WHPB-MeOH can be found in a companion paper [7]. The samples for WHPB and its MeOH extract were diluted to the designated concentration using filtered artificial seawater (Table 1). The maximum DMSO in the test solutions did not exceed 0.1%.

2.2. Biological Materials

The test species—marine copepod P. parvus sl—was collected between October and December 2017 from the coast of Geoje city in South Korea. Zooplankton were captured using a conical net (0.42 m mouth diameter, 200 μm mesh size) towed vertically from the bottom to the surface at 0.5 m s⁻¹. The zooplankton thus collected were diluted with ambient sea water to 5 L and placed in cooler boxes during the return to the laboratory. Ambient seawater for subsequent incubation was also taken from approximately 1 m below the surface using a 5 L Niskin-type sampler and transferred into pre-cleaned 2 L polyethylene bottles. The samples were then kept at 4 °C and brought to the laboratory for analysis. YSI 6600 V2 Sonde (YSI Inc., Yellow Springs, OH, USA) cast from the near-bottom to the surface recorded the in situ water quality, including temperature, salinity, and dissolved oxygen. Ambient seawater was filtered through a 47 mm diameter Whatman GF/F glass-fiber filter (Whatman International Ltd., Maidstone, UK) and sterilized at 121 °C for 15 min using an autoclave. From the collected zooplankton samples, 15 individual females of healthy adult P. parvus sl were segregated using a stereomicroscope and stored in petri dishes containing 20 mL filtered and sterilized seawater for 6 h in an incubator set to in situ water temperature. All eggs produced during incubation were collected using a 50 μm mesh, and 10 eggs were placed in a sterile 6-well culture flask with 10 mL of each toxicant solution, together with control (DMSO 0%) and negative control (DMSO 0.1%). All experiments were performed at in situ temperatures with a 12 h light–dark cycle and illumination intensity of 60 μmol m⁻² s⁻¹. Each treatment was replicated four times. Hatching and nauplius immobilization were observed under a stereomicroscope after 24 and 48 h. The egg-hatching rate was calculated as the ratio of the number of hatched embryos versus total eggs inoculated. Furthermore, the nauplius mortality was calculated as the ratio of the number of immobilized or dead embryos to hatched embryos. Immobilized nauplii (which did not move vertically and/or horizontally in samples) were observed during 15 s and confirmed through stimulation with a needle.

2.3. Statistical Analysis

The results were presented as means and standard deviations of the raw data. The significance of differences between the experimental groups was assessed through a t-test (SPSS, Inc., IBM, Chicago, IL, USA). The effects of three booster biocides and two WHPB samples on the marine copepod P. parvus sl were evaluated through LC₅₀ calculation using R studio (R Studio, Inc., Boston, MA, USA). Statistical significance was estimated at a 95% confidence interval (95% CI) to confirm the overlap of the LC₅₀ values.

3. Results and Discussions

3.1. Effect of DMSO on P. parvus sl Egg-Hatching Rate and Nauplius Mortality

During the collection of P. parvus sl, the mean ± standard deviations of in situ water temperature (℃), salinity (psu), DO (mg L⁻¹), and DO sat (%) were 23.3 ± 1.7℃, 29.6 ± 0.8, 6.8 ± 0.6 mg L⁻¹, and 95.3 ± 8.3%, respectively. Before the experiment, we pre-checked the effect of DMSO on the egg-hatching rate and nauplius mortality of P. parvus sl. We found no difference between control (DMSO 0%) and negative control (DMSO 0.1%), with mean + standard deviations of 91.1 ± 2.8% and 89.3 ± 3.8% in egg-hatching rate and 9.4 ± 3.0% and 8.6 ± 0.8% in nauplius mortality (t-test, p = 0.88), respectively. Furthermore, the mean ± standard deviations of pH and DO sat (%) in all control and treatment experiments were 8.04 ± 0.03 and 98.38 ± 0.37, respectively, and no effects of other environmental factors except for the toxic test substances were found during the culture process.

3.2. Toxicity Test of Antifouling Biocides

The egg-hatching success of P. parvus sl exposed to various concentrations of three biocides up to 100 μg L⁻¹ for 24 h was inhibited at the lowest concentration of each test biocide (Figure 1a–c). In particular, the concentration of three booster biocides, under which the egg-hatching rate of P. parvus sl decreased below 50% compared to the control treatment, was 5 mg L⁻¹ for Diuron and Irgarol 1051 and 0.01 mg L⁻¹ for Sea-nine 211. Thus, Sea-nine 211 showed the most adverse effect on the egg-hatching rate among the three booster biocides. Diuron and Irgarol 1051 exhibited similar tendencies until the egg-hatching rate reached 50%. However, as the concentration increased further, the egg-hatching rate in groups treated with Diuron decreased to zero, whereas the egg-hatching rate in the groups treated with Irgarol 1051 were 25% at the highest concentration (15 mg L⁻¹), indicating that Irgarol 1051 was less effective than Diuron at concentrations between 2.0 and 15 mg L⁻¹. The mortality rate of nauplii exposed to various concentrations of Diuron and Irgarol 1051 was 45% and 60%, respectively, at 2 mg L⁻¹ and 100% at 7.5 mg L⁻¹ (Figure 1d–f). Moreover, the mortality rate of nauplii exposed to Sea-nine 211 was 61.5% at 0.001 mg L⁻¹ and 100% at 0.01 mg L⁻¹. Therefore, the median lethal concentrations after 24 h (24 h LC₅₀) at 95% confidence interval were similar for Diuron and Irgarol 1051 at 1.968 and 1.391 mg L⁻¹, respectively, while the 24 h LC₅₀ of Sea-nine 211 was 0.438 μg L⁻¹, which indicated approximately 3000 times greater toxicity compared with Diuron and Irgarol 1051 (Figure 1,

Table 2. LC₅₀ values (LC₅₀ ± 95% CI) of Paracalanus parvus sl at 24 h for three booster biocides (Diuron, Irgarol 1051, and Sea-nine 211) and wastewater (WHPB and WHPB-MeOH) from high-pressure water blasting.

Material	Unit	95% CI				t-Value	p-Value
		24-h LC50	Std. Error	Low Limit	High Limit
Diuron	μg L−1	1968	136	1832	2104	14.4505	1.1 × 10−11
Irgarol 1051		1391	276	1115	1667	5.0481	4.7 × 10−5
Sea-nine 211		0.438	0.096	0.342	0.534	4.5443	2.2 × 10−4
WHPB (1)	Dilution factor	11,820	2459	9361	14,279	48.0738	2.2 × 10−16
WHPB-MeOH (2)		1079	145	934	1224	7.4148	2.0 × 10−7

⁽¹⁾ WHPB: Wastewater from high-pressure water blasting. ⁽²⁾ WHPB-MeOH: WHPB-MeOH extract

). Both Diuron and Irgarol 1051 are herbicide-based biocides that inhibit photosynthesis by blocking electron transport in chloroplasts [25,26]. Because of this mechanism, these biocides present a higher toxicity for macroalgae and phytoplankton species than marine invertebrates [13,17]. Similarly, Sea-nine 211 is more fatal to animal cells compared with the other two compounds and is known to cause necrosis via oxidative stress in animal cells [27,28]. Moreover, previous bioassay experiments using Daphnia magna, Vibrio fischeri, and cultured fish cell suspensions have shown that Sea-nine 211 is more toxic than Irgarol 1051 and Diuron, e.g., [29]. Jung et al. (2017) [13] reported the 48 h LC₅₀ of Diuron, Irgarol 1051, and Sea-nine 211 in Artemia larvae as 30,573, 9734, and 318 μg L⁻¹, respectively, indicating the greater toxicity of Sea-nine 211 compared with Diuron and Irgarol 1051.

From the previous study on the concentration of Sea-nine 211 in seawater, which was found to be the most toxic in the present study, the concentration of Sea-nine 211 was 0–6 ng L⁻¹ in Jinhae Bay, Korea [10]; <0.3–4 ng L⁻¹ in Osaka port, Japan [30]; <6.3–49 ng L⁻¹ in Patras marina in Greece; and <5–283 ng L⁻¹ at Denmark’s harbors [31], which showed little difference but was detected at a low concentration. However, a higher concentration of 2.6–3.7 μg L⁻¹ [32,33] was detected at Catalonia marina, Spain, which was approximately five times higher than the 24 h LC₅₀ for P. parvus sl nauplii. This may affect the growth and mortality of natural copepod nauplii. In addition, the majority of copepods are primary consumers of phytoplankton; therefore, research on the indirect effect on the growth and reproduction of copepods feeding on phytoplankton exposed to biocides such as Diuron and Irgarol 1051 is needed. Moreover, although the Sea-nine 211 concentrations found in previous studies have been considered too low to cause acute toxicity in marine animals, determining their chronic toxic effects is necessary.

The sensitivity of P. parvus sl to the three booster biocides was compared with the results of previous studies to determine its suitability as a toxicity indicator of antifouling booster biocides. In the case of Daphnia magna, Diuron 48 h EC₅₀ and Sea-nine 24 h EC₅₀ concentrations were 8600 and 8 μg L⁻¹, respectively [34]. The 24 h LC₅₀ concentration of the benthic copepod Tigriopus japonicus nauplii was >4000 μg L⁻¹ in Irgarol 1051 [9,35] and 23 and 77 μg L⁻¹ in Sea-nine 211 [36,37], indicating that both species were less sensitive than P. parvus nauplii were in the present study. Furthermore, the 24 h LC₅₀ concentrations of Artemia salina larvae were reported to be 23,270 and 12,500 μg L⁻¹ for Diuron and >4000 μg L⁻¹ for Irgarol 1051 [38,39], which was higher than the 24 h LC₅₀ concentration of P. parvus sl nauplii in the present study. In the case of Sea-nine 211 study results, the 24 h LC₅₀ concentrations of Portunus trituberculatus larvae and Amphibalanus amphitrite larvae were >101 and 340 μg L⁻¹, respectively [36,40]. These were approximately 200 times higher than the result of the present study, which was confirmed to be 0.438 μg L⁻¹. Furthermore, embryos of the mollusk Mytilus edulis were also less sensitive than P. parvus sl nauplii in Sea-nine 211 exposure experiments (

Table 3. LC₅₀ and EC₅₀ concentration of three booster biocides (Diuron, Irgarol 1051, and Sea-nine 211) on zooplankton species.

Biocide	Phylum, Order	Species	Effect Level	Concentration	Reference
				(µg L−1)
Diuron	Arthropoda, Cladocera	Daphnia magna	48 h EC50	8600	[34]
	Arthropoda, Anostraca	Artemia Larvae	48 h LC50	30,573	[13]
		Artemia salina (Larvae)	24 h LC50	23,270	[38]
			24 h LC50	12,500	[39]
	Arthropoda, Decapoda	Palaemon serratus (Larvae)	24 h EC50	3011	[17]
	Arthropoda, Amphipoda	Hyalella azteca	96 h LC50	19,400	[34]
Irgarol 1051	Arthropoda, Cladocera	Daphnia magna	24 h LC50	16,000	[35]
		Daphnia pulex	24 h LC50	5700	[35]
	Arthropoda, Anostraca	Artemia salina	24 h LC50	>4000	[35]
	Arthropoda, Harpacticoida	Tigriopus japonicus (Larvae)	24 h LC50	>4000	[9]
Sea-nine 211	Arthropoda, Cladocera	Daphnia magna	24 h EC50	8	[34]
	Arthropoda, Anostraca	Artemia salina (Larvae)	48 h LC50	318	[13]
	Arthropoda, Calanoida	Acartia tonsa	48 h LC50	16.1	[15]
	Arthropoda, Harpacticoida	Tigriopus japonicus (Larvae)	24 h LC50	23	[36]
			24 h LC50	77	[37]
	Arthropoda, Decapoda	Portunus trituberculatus (Larvae)	24 h LC50	>101	[36]
	Arthropoda, Sessilia	Amphibalanus amphitrite (Larvae)	24 h LC50	340	[40]
	Mollusca	Mytilus edulis (embryonic)	24 h LC50	11	[18]

). The results of comparison between the present study and previous studies showed that P. parvus sl nauplii exhibited higher sensitivity to antifouling biocides compared with other zooplankton species. Therefore, considering the status of P. parvus sl nauplii in the marine ecosystem food web, they could be used as an indicator for evaluating developmental toxicity to antifouling paint in the marine environment.

3.3. Concentration of Metal and Organic Biocides

The materials in the effluents generated when the ship’s hull surface is cleaned using high-pressure water jets are generally categorized as heavy metals and booster biocides. Among heavy metals in the effluents, the concentrations of zinc and copper were the highest, at 12,269 ± 473 and 6964 ± 129 μg L⁻¹, respectively, accounting for more than 93% of the total heavy metal concentration, which was followed by Ba, Mn, Fe, and As. The concentrations of metal elements detected in the WHPB samples were compared with the marine ecosystem protection standards included in Korean marine environment standards (

Table 4. Comparison of dissolved metal concentration in WHPB samples with marine environmental standards for metals in Korea.

		Dissolved Metals (μg L−1, ppb)
		Cr	Mn	Fe	Co	Ni	Cu	Zn	As	Se	Cd	Ba	Pb
WHPB (1)	Mean	0.92	439	260	1.53	7.28	6964	12,269	53	3.52	1.88	528	0.80
	Std	0.17	12	82	0.16	0.13	129	473	1	1.32	0.08	13	0.14
KMEPC (2)	Mean	200	-	-	-	11	3	34	9.4	-	19	-	7.6

⁽¹⁾ WHPB: Wastewater from high-pressure water blasting. ⁽²⁾ KMEPC: Marine Ecosystem Protection Criteria in marine environmental standards; criteria for comparison with one-time observations.

). The copper and dissolved zinc concentrations were found to be approximately 2300 and 360 times higher, respectively, than the Korean marine environment standards. Arsenic was also approximately five times higher than the marine ecosystem protection standard. These results indicate that if effluents are released from the hull surface during in-water cleaning without a capture process, they could have detrimental effects on marine organisms.

In the booster biocide, zinc pyrithione (ZnPT) concentration was approximately 259.87 μg L⁻¹, which was higher than that of copper pyrithione (CuPT) (104.61 μg L⁻¹). The antifouling paint applied to the research vessels was confirmed to contain more zinc-based metallic biocides than copper. However, booster biocides such as Diuron, Irgarol, and Sea-nine were not detected.

3.4. Toxicity Test of WHPB and WHPB-MeOH

The eggs of P. parvus sl in the WHPB experiment group did not hatch in either 100- or 50-fold dilutions and were 17.0 ± 6.8% in 1000-fold dilution, 23.3 ± 11.5% in 5000-fold dilution, 30.0 ± 0.0% in 7500-fold dilution, 53.3 ± 11.5% in 8000-fold dilution, and 86.7 ± 5.8% in 10,000-fold dilution (Figure 2a). In the WHPB-MeOH experimental groups, the egg-hatching success rates of P. parvus sl were inhibited up to 110-fold dilution and were 23.3 ± 5.8% in 165-fold dilution, 33.3 ± 5.8% in 220-fold dilution, and 33.3 ± 5.8% in 1099-fold dilution (Figure 2b). Furthermore, the mortality of P. parvus sl in the WHPB experiment groups was 10.3 ± 0.0% in control experiments and 15.4 ± 5.8% at 10,000-fold dilution, 53.3 ± 11.5% in 8000-fold dilution, and 88.9 ± 5.8% in 7500-fold dilution in dilution experiments. No living nauplii were found in the 5000-, 1000-, 100-, and 50-fold dilution experiments (Figure 2c). In the case of the WHPB-MeOH experimental groups, the mortality of P. parvus sl was 10.3 ± 0.0% in control experiments and 11.5 ± 5.8% in 2198-fold, 42.9 ± 10.0% in 1099-fold, and 90.0 ± 10.0% in 220-fold dilutions in dilution experiments. Moreover, no living nauplii were found in the 110-, 22-, 2-, and 1-fold dilution experiments (Figure 2d). Thus, the WHPB experimental groups had a more negative effect on the egg-hatching rate and nauplius mortality of P. parvus sl, although the WHPB-MeOH experimental groups also showed high toxicity. This could be confirmed more clearly from the LC₅₀ results, which showed approximately 11,820-fold dilution in WHPB and approximately 1079-fold dilution in WHPB-MeOH (Table 2). Thus, the WHPB containing heavy metals such as copper and zinc was confirmed to have a dilution factor approximately 13 times higher than that of WHPB-MeOH. The expected metallic-element concentration at 24 h LC₅₀ in WHPB was 0.59 μg L⁻¹ for copper and 1.04 μg L⁻¹ for zinc, and the booster biocide concentration was 0.022 and 0.009 μg L⁻¹ for ZnPT and CuPT, respectively (Figure 3). Furthermore, the expected concentration of organic biocide at 24 h LC₅₀ in WHPB-MeOH was 0.24 and 0.009 μg L⁻¹ for ZnPT and CuPT, respectively (Figure 3).

We analyzed the toxic effects of WHPB relative to the three booster biocides, which were prohibited or banned in some European countries [41,42]. As described in the previous paragraph, the 24 h LC₅₀ concentrations of Diuron, Irgarol 1051, and Sea-nine 211 affecting hatched larvae were 1.97, 1.39, and 0.0004 mg L⁻¹, respectively. By contrast, the 24 h LC₅₀ concentrations of ZnPT and CuPT in WHPB affecting the egg-hatching rate were found to be 0.02 and 0.01 μg L⁻¹, respectively. This result confirms that WHPB toxicity for the egg-hatching rate and nauplius mortality of P. parvus sl is stronger for Sea-nine 211, which was 3000 times more toxic than Irgarol 1051 and Diuron in the present study.

4. Conclusions

In this study, we aimed to determine the acute toxicity effects on the egg-hatching rate and nauplius mortality of P. parvus sl exposed to three types of booster biocides and WHPB. Sea-nine 211 was confirmed to be more toxic than Irgarol 1051 and Diuron. However, even Sea-nine 211 was found to have approximately 20 times lower acute toxicity effects than WHPB (ZnPT-based antifouling paint) obtained using a brush from the ship’s hull surface. Therefore, assuming that the entire hull surface of a container ship with a length of more than 200 m is cleaned, hundreds of tons of wastewater would be discharged into the port water, which is expected to profoundly impact the port environment and ecosystem, considering the results of this study that WHPB affects the egg-hatching rate and nauplius mortality of P. parvus sl even when diluted more than 1000 times. Furthermore, wastewater contains a large number of antifouling paint particles, which are deposited on the bottom layer and continuously leak pollutants into the harbor over a long period of time during their decomposition. Moreover, most booster biocides including ZnPT are generally understood to transform into more toxic by-products via biodegradation and photolysis. In conclusion, if wastewater generated during in-water cleaning activities is directly discharged into port water and coastal areas without being captured by a post-treatment system, it is expected to negatively affect the functioning and structure of the microbial food web in the port water.