Sediment samples
The concentrations of TBT in sediment samples from Harima Nada Sea and Osaka Bay were in the range of 5.5 - 17 ng/g dw and 2.0 - 28 ng/g dw, respectively (Table 5). The concentrations of TBT in sediment samples were in the ranged of <0.3 – 1, 980 ng/g dw after the worldwide ban as shown as table 4. The concentrations of TBT in the sediment collected from Harima Nada Sea and Osaka Bay were lower than those reported values. However, it has been reported that background concentrations in sediment from a French aquatic environment ranged from <0.25 - 1.16 ng/g dw (Cavalheiro et al. 2016). Tributhyltin values measured in this study were higher than the background levels in France. On the other hand, the concentrations of TBT in sediment samples from the Port of Osaka before the worldwide TBT ban ranged from 10 – 2,100 ng/g dw, (Harino et al.1998), indicating that the input of TBT had decreased by the effect of AFS convention.
The concentrations of ΣBTs in sediment samples are also shown in Table 5. The concentrations of ΣBTs in Harima Nada Sea and Osaka Bay ranged from 44 – 1,900 ng/g dw and 200 – 370 ng/ng dw, respectively. When the concentrations of TBT detected in sediment samples from Harima Nada Sea were compared with those detected in sediment samples from Osaka Bay, no drastic difference in TBT concentration was observed between the two regions. The reported concentrations ofΣBTs in sediment samples ranged from 7.9 – 2,120 ng/g dw after the worldwide ban. Therefore, the ΣBTs concentrations in sediment samples from Harima Nada Sea and Osaka Bay were similar to the concentrations previously reported in other areas.
The horizontal distribution of ΣBTs was investigated in Harima Nada Sea and Osaka Bay. The concentrations of ΣBTs were high at St.H4 in Harima Nada Sea and Sts.O1-O2 in Osaka Bay. It is considered that these sampling stations were once heavily contaminated by OT compounds. Moreover the water depth at St.H4 was 40 m, which is greater than the depth at other stations. It was reported that the degradation of TBT is slower under anaerobic conditions (Dowson et al.1993; de Mora et al.1989). In addition, it is considered that many fishing boats sailed in the vicinity of this station, which would lead to a high load of TBT, because fishing is the main industry around St.H4. Station O1 is located at the mouth of a large river called the Yodo River. Because this area is used as an international trading port, there are many factories around St. O1. Industry is also the main activity around St.O2, which is surrounded by trading ports and international airports. Furthermore, it is considered that chemical pollutants flow in a direction from St.O1 to St.O2, since this is the direction of water flow in Osaka Bay. Therefore, it is probable that the concentrations of ΣBTs were high at these sampling sites.
The concentrations of TPT in sediment samples from Harima Nada Sea and Osaka Bay were in the range of <0.1 - 2,700 ng/g dw and 0.2 - 1,300 ng/g dw, respectively (Table 5). Although there are few reports on TPT concentrations after the worldwide ban, there values were ranged <0.5 – 346 ng/g dw (Table 4). Judging from the reported values, the concentrations of TPT in Harima Nada Sea and Osaka Bay were higher than the previously reported values.
The concentrations of ΣPTs in sediment samples from Harima Nada Sea and Osaka Bay ranged from 8.3 - 2,700 ng/g dw and 6.0 - 1,300 ng/g dw, respectively. Few recent papers have measured all three of MPT, DPT and TPT. It was reported thatΣPTs were not below the detection limit in most samples from seaports on the Gulf of Gdansk on the southern Baltic coast, and the maximum concentration ofΣPTs in the samples was 660 ng/g dw (Table 4), indicating that the concentrations of ΣPTs in Harima Nada Sea and Osaka Bay were fairly high.
Next, we will consider the horizontal distribution ofΣPTs. The highest concentrations ofΣPTs were observed at. St.H8 in Harima Nada Sea and Sts.O1 - O2 in Osaka Bay, the main industries at St. H8, are tourism and fishing. Higher concentrations ofΣPTs in sediment were observed not only in the industrial and fishing areas but also in the tourist areas of all three stations. the horizontal distribution of TPT was similar to that ΣPTs.
With regard to BT compounds, the horizontal distributions of ΣBTs and TBT were different, but in the case of PT compounds, the horizontal distributions of TPT were similar to those of ΣPTs. This may be due to the difference in decomposition of TBT and TPT in sediment. In addition about half of the PDI values at each station were 1 or less. This shows that there was new input of TPT to sediment. As to the possible reasons, it may be that TPT was used as a pesticide.
Partition coefficients between water and sediment
The partition coefficients between water and sediment (Ksw) of each BT were calculated by dividing the concentration of each BT in the sediment by the concentration of each BT in the bottom water (Fig.4). The Ksw values of MBT, DBT and TBT were 0.9 X 104 - 2 X 104, 2 X 104 - 56 X 104 and 0.2 X 104 - 26 X 104, respectively. Harino et al. (1998) reported that Ksw of TBT 3.8 X 104 in the Port of Osaka before BTs ban by IMO. Furthermore, the Ksw of each BT was higher in the order TBT > DBT > MBT. These results in this study were similar to those reported before the ban. Although the concentration in water and sediment samples was lower than those before the ban, the Ksw values were similar, suggesting that OTs were eluted from sediment to water. The Ksw of PTs could not be calculated, because PT was not detected in the bottom water.
Alternative biocides
Seawater samples
The concentrations of each alternative biocide in surface water samples were compared between Harima Nada Sea and Osaka Bay, and their levels were compared to the reported values (Table 5). In addition, the horizontal distributions of each alternative biocide in Harima Nada Sea and Osaka Bay and the effect of antifouling biocide for aquatic organisms were discussed.
Diuron in surface water samples from Harima Nada Sea and Osaka Bay was detected in the ranges of <1 - 5.6 ng/L and <1 - 53 ng/L, respectively and the levels of diuron in Harima Nada Sea were lower than those in Osaka Bay. The concentrations of diuorn in the worldwide were ranged in <0.1 – 70 ng/L as showing in Table 7. The levels of diuron in Harima Nada Sea and Osaka Bay were lower than levels those in the other sampling stations. These values were compared with those in water samples from the Port of Osaka before the ban of diuron by IMO. The concentration of diuron was in the ranged of 0.8 - 267 ng/L (Harino et al. 2005) .The concentrations of diuron in this study were lower than those before the ban.
The concentrations of diuron that were detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. It was reported that the values of LC50 for Crustacea and fish were over 3,000 mg/L and 7,800 mg/L, respectively (Bao et al. 2011). Nebeker and Schuytema (1998) reported that the NOAEL values for freshwater cladocerans, amphipods, midges, minnows, worms, and snails ranged from 1,800 - 20,000 mg/L. Onduka et al. (2022) reported that NOEC for pacific oyster (Crassostrea gigas) embryos was 9,500 ng/L. The concentrations of diuron in water samples from Harima Nada Sea and Osaka Bay were lower than the values associated with toxicity level for aquatic organisms
Sea-Nine 211 in surface water samples from Harima Nada Sea was detected in the range of <1 - 1.8 ng/L, while Sea-Nine 211 was not detected in water samples from Osaka Bay. There have been few reports concerning the concentrations of Sea-Nine 211 (Table 7). The concentrations of these compounds in the other aquatic areas ranged from <0.1 – 31 ng/L. The levels of Sea-Nine 211 in Harima Nada were lower than those in other aquatic areas. Next, we compared these values with the concentration of Sea-Nine in water from the Port of Osaka before the ban of TBT by IMO. Before the ban, the concentrations of Sea-Nine 211 were in the range of <3 - 4 ng/L (Harino et al. 2005) .The concentrations of Sea-Nine 211 in this study were thus lower than those before the ban
The concentrations of Sea-Nine 211 that were detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. It has been reported that the LC50 for Crustacea were over 0.42 – 4 mg/L within 72 hrs. (Femandez-Alba et al. 2002 ; Myers et al. 2006). Onduka et al. (2022) reported that NOEC for pacific oyster (Crassostrea gigas) embryos was <3 ng/L . Therefore, the concentrations of Sea-Nine 211 detected in these sea regions were lower than LC50 values. The chronically toxic levels of Sea-Nine 211 for crustaceans (Daphnia magna) and fish (Onorhynchus tshawytcha) has been shown to range from 1,200 ng/L to 14,000 ng/L (Shade et al, 1994: Okamura et al 2002). Mochida et al (2010) reported that 96-h LC50 values for Red sea bream and mummichog were 5,100 and 4,700 ng/L. It is reported that the lowest 72-h EC50 of Sea-Nine 211 was 160 ng/L for the alga (Skeletonema costatum) and directly disrupts cell division or biosynthetic pathways related to cell devision (Mochida et al., 2015). Onduka et al (2013) reported that 72h NOEC of Sea-Nine 211 for Dunaliella tertiolecta, Tetraselmis tetrathele, Chaetoceros calcitrans, Skeletonema Costatum which are belong to algae, were 1200, 1700, 480, 40-120 and 60-70 ng/L, respectively, and 24-h EC50 for Tigriopus japonicas and Portunus trituberculantus which are belong to crustacea were 1400 – 1800 ng/L and 11,000 – 13,000 ng/L, respectively. The concentrations of Sea-Nine 211 which were detected in this study, were lower than the toxic level to aquatic organisms.
Irgarol 1051 in water samples from Harima Nada Sea and Osaka Bay was detected in the ranges of <1 - 2.5 ng/L and <1 - 4.0 ng/L, respectively. The reported concentrations of Irgarol 1051 ranged from <0.1 – 55 ng/L (Table 7). The concentrations of Irgarol 1051 measured in this study were thus lower than the reported values. These values were compared with those in water samples from the Port of Osaka before the ban of TBT by IMO. The concentration of Irgarol 1051 ranged from <0.8 - 267 ng/L (Harino et al. 2005). The concentrations of Irgarol 1051 in this study were lower than those before the ban.
The concentrations of Irgarol 1051 detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. Mochida et al. (2019) researched the physiological effects of eelgrass (Zostera marina) by Irgarol 1051. As the result, the toxic effect on growth was observed at the concentrations of Irgarol 1051 of 1,000 ng/L and higher. It is has been reported that the LC50 values for Crustacea and fish were over 5,700 mg/L (Okamura et al. 2000, 2002; Toth et al. 1996; Bao et al. 2011). Onduka et al. (2022) reported that NOEC for pacific oyster (Crassostrea gigas) embryos was 7,000 ng/L. The concentrations of Irgarol 1051 were also compared to the endpoint of chronic toxicity to the growth and survival of crustacean (Mysidopsis bahia) and fish (Oncorhynchus mykiss). The endpoint values ranged from 4,020 - 110,000 ng/L (Hall Jr et al.1999). Judging from the concentration of Irgarol 1501 detected in Harima Nada Sea and Osaka Bay, we considered that there was very low possibility that the Irgarol 1051 would affect the aquatic organisms in these regions.
The concentrations of M1 which is the degradation product of Irgarol 1051were in the range of 1.9 – 33 ng/L. The ratio of M1 to Irgarol 1051 was calculated as being range of 1.2 – 34, suggesting that the decomposition rate of Irgarol 1051 exceeded its input rate. The reported concentrations of M1 ranged from <3.2 - 63.4 ng/L (Table 7). The concentrations of M1 in this study were similar to those observed in other aquatic areas. These values were compared with those in water samples from the Port of Osaka before the ban of M1 by IMO. The concentrations of M1 were in the range of <1.9 - 167 ng/L (Harino et al. 1998). The concentrations of M1 in this study were lower than those before the ban.
The concentrations of M1 that were detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. It has been reported that 96hr LC50 concentrations for growth of microalgae were 73 - 83 mg/L of 96hr (Gatidou and Thomaidis 2007). The levels of M1 that were detected in these sea areas were lower than the LC50 values.
Dichlofluanid in water samples from Harima Nada Sea and Osaka Bay was detected in the ranges of 28 - 61 ng/L and <1 – 343 ng/L, respectively. The concentrations of dichlofluanid in Osaka Bay were higher than those in Harima Nada Sea. There have been few reports concerning of the concentrations of dichlofluanid (Harino 2016). Harino and Yamato (2021) reported that dichlofluanid concentrations of <0.1 – 44 ng/L in Tanabe Bay. The concentrations of dichlofluanid measured in the present analysis were thus higher than those reported previously.
The concentrations of dichlofluanid that were detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. It was reported that the values of 96 hr LC50 values Crustacea were 133 – 1,050 mg/L (Fernandez-Alba et al. 2002). Bellas (2006) reported that the 48 hr LC50 values for mussels and sea urchins were 627 mg/L and 81 mg/L, respectively. The levels of dichlofluanid in Harima Nada Sea and Osaka Bay were lower than the LC50 values for aquatic animals. The NOAEL for fish and invertebrates of aquatic organisms ranged from 2.65 – 4.55 mg/L (UK 2016). The concentrations of dichlofluanid were lower than the values causing acute toxicity and lower than the NOAEL.
Chlorothalonil in water samples from Harima Nada was not detected. On the other hand, chlorotharonil was detected at only St.O2 in Osaka Bay, and its concentration was near the detection limit. Harino and Yamato (2021) reported that the chlorothalonil concentrations ranged from 8 - 26 ng/L in Tanabe Bay as shown in table 7. The concentrations of chlorothalonil in Harima Nada Sea and Osaka Bay were lower than those in Tanabe Bay. The high concentrations of chlorothalonil in Tanabe Bay may be due to the use as a pesticide, because Tanabe Bay is surrounded by forest areas.
The concentrations of chlorothalonil that were detected in Harima Nada Sea and Osaka Bay were compared to the acute toxicity levels for aquatic organisms. It was previously reported that the 96hr LC50 values for Crustacea were 69 - 110 mg/L (Bao 2011). Bellas (2006) reported that the 48 hr LC50 values for mussels and sea urchins were 8.7 mg/L and 6.6 mg/L, respectively. 72h EC50 and 72 hr NOEC of Chlorothalonil for alga (Skeletonema Costatum) were 950 ng/L and 560 ng/L. 24 hr EC50 of crustacea (Tigriopus japonicus) was 16,000 ng/L 96 hr LC50 of Kuruma prawn (Marsupenaeus. japonicas), red sea bream (Pagrus major) and mummichog (Fundulus heeroclitus) were 290,000 ng/L, 35,000 ng/L and 61,000 ng/L, respectively and 8-week NOEC and LOEC for mummichog was 11,000 ng/L and 32,000 ng/L. It has been reported that the NOAELs of chlorothalonil for fish and invertebrates are 1,300 and 600 ng/L (Thistle and Durkin 2015). Onzuka et al. (2012) reported that in the early-live-stage test with mummichog empryos, the lowest- and no-observed-effect concentrations, based on the measured concentrations, were 32,000 and 11,000 ng/L, respectively. The concentrations of chlorothalonil in Harima Nada Sea and Osaka Bay were lower than the toxicity levels for aquatic organisms.
The concentration of each alternative biocide was compared between surface water and bottom water (Fig.3). The concentration of diuron, Sea-Nine 211 and Irgarol 1051 in the surface water samples were higher than those in the bottom water samples. The concentrations of dichlofluanid in the surface water samples were also higher than those in the bottom water samples in 6 of 11 sampling sites. Alternatives biocides showed the opposite trend compared to OTs. Collectively, indicating that the detected alternative biocides were not re-eluted from the sediment, but rather eluted from the ship hull.
The above results can be summarized as follows. The concentrations of antifouling agents detected in Harima Nada Sea and Osaka Bay were lower than those detected in other areas, except for dichlorofluanid. In terms of the reason why this compounds was detected, we considered that dichlorofluanid may have been used as a pesticide in addition to its use as an antifouling paint on the ship hulls. More specifically, the concentrations of dichlorofluaind may have been higher than in other aquatic regions because greater amounts of dichlorofluanid were used as a pesticide around the aquatic areas. Furthermore, it found that the concentrations of alternative biocides in Harima Nada Sea and Osaka Bay are not at a level that affects aquatic organisms.
Sediment samples
The concentrations of diuron in sediment ranged from 32 - 488 ng/g dw (mean 117 ng/g dw) and 88 - 342 ng/g dw (mean 153 ng/g dw) a in Harima Nada Sea and Osaka Bay, respectively (Table 8). The concentrations of diuron in Harima Nada Sea and Osaka Bay were similar. These values were compared with the reported values. The reported concentrations of diuron in sediment samples ranged from 0.01 – 1, 112 ng/g dw. The concentrations of diuron in Harima Nada Sea and Osaka Bay were within the reported values. These values were compared with those in sediment samples from the Port of Osaka before the worldwide diuron ban. The concentration of TBT was in the ranged of 10 – 2,100 ng/g dw, (Harino et al. 2005) .The concentrations of diuron in this study were thus lower than those before the ban of TBT.
The concentrations of Sea-Nine 211 in sediment ranged from 47 – 591 ng/g dw (mean 135 ng/g dw) and 63-93 ng/g dw (mean 75 ng/g dw) a in Harima Nada Sea and Osaka Bay, respectively. The concentrations of Sea-Nine 211 in Harima Nada Sea were higher than those in Osaka Bay. There are few reports on the concentrations of Sea-Nine 211 in sediment samples (Table 9). It was reported that the concentrations of Sea-Nine 211 were in the range of <038 – 81.6 ng/g dw in sediment from Panama (Batista-Andrade et al. 2018). Harino and Yamato (2021) reported that Sea-Nine 211 like diuron, was not detected in Tanabe Bay. The concentrations of Sea-Nine 211 in Harima Nada Sea and Osaka Bay were higher than the reported values. These values were compared with those in sediment samples from the Port of Osaka before TBT ban by IMO. The concentrations of Sea-Nine 211 were in the range of 10 – 2,100 ng/g dw, (Harino et al.1998) .The concentrations of Sea-Nine 211 in this study were lower than those before the ban. Onduka et al. (2013) reported that 14 d LC50 and 14 d NOEC of growth 110 ng/g dw and 9.7 ng/g dw and the lowest value of 72 hr NOEC was 0.04 ng/g dw for the alga (Chaetoceros calcitrans).. Because the concentrations of Sea-Nine 211 were higher than NOEC for alga, there is concern about the impact on benthic organisms.
The concentrations of Irgarol 1051 ranged from 33 – 128 ng/g dw (mean 57 ng/g dw) and 43-83 ng/g dw (mean 56 ng/g dw)in Harima Nada Sea and Osaka Bay, respectively. The concentrations of Irgarol 1051 in Harima Nada Sea were similar to those in Osaka Bay. The reported values of Irgarol 1051 ranged from <0.08 – 1,112 ng/g dw (Table 9). Although the concentrations of Irgarol 1051 in Harima Nada Sea and Osaka Bay were lower than those in the Seto Inland Sea, the concentrations of Irgarol 1051 in Harima Nada Sea and Osaka Bay were higher than those in the other areas. These values were compared with those in sediment samples from the Port of Osaka before the TBT ban by IMO, in the Port of Osaka, the concentration of Irgarol 1051 range from 10 – 2,100 ng/g dw, (Harino et al.1998) .The concentrations of Irgarol 1051 in this study were thus lower than those the before ban.
The concentrations of M1 ranged from 60 - 128 ng/g dw (mean 137 ng/g dw) and 104 – 377 ng/g dw (mean 174 ng/g dw in Harima Nada Sea and Osaka Bay. Although the concentrations of Irgarol 1051 in Harima Nada Sea were similar to those in Osaka Bay, The concentrations of M1 in Osaka Bay were higher than those in Harima Nada Sea. It is reported that M1 was not detected in Tanabe Bay (Harino and Yamato 2021). The detection of M1 at relatively high concentrations is noteworthy at present.
The concentrations of dichlofluanid ranged from 67 - 8,038 ng/g dw (mean 1,130 ng/g dw) and 104 -263 ng/g dw (mean 162 ng/g dw in Harima Nada Sea and Osaka Bay, respectively. The concentrations of dichlofluanid in Harima Nada Sea were higher than those in Osaka Bay. It was reported that dichlofluanid was not detected in sediment from Panama (Batista-Andrade et al. 2018). Harino and Yamato (2021) reported that dichlofluanid was also not detected in Tanabe Bay (Table 9). Therefore, the concentrations of dichlofluanid in Harima Nada Sea and Osaka Bay were higher than those in Tanebe Bay, suggesting the use of chlorothalonil used as a pesticide in past tense.
The concentrations of chlorothalonil ranged from 31 - 2,975 ng/g dw (mean 428 ng/g dw) and 49 -165 ng/g dw (mean 83 ng/g dw) in Harima Nada Sea and Osaka Bay, respectively. The concentrations of chlorohalonil in Harima Nada Sea were higher than those in Osaka Bay. There have been only a few studies on chlorothalonil. Harino and Yamato (2021) reported that the concentrations of chlorothalonil ranged from <0.1 – 8.2 ng/g dw in Tanabe Bay. These results demonstrated that the concentrations of alternative biocides in sediment from Harima Nada Sea and Osaka Bay were higher than those in Tanabe Bay. Similar to the water samples, it may be the discharge of chlorothalonil used as a pesticide to the sea areas.
The above results can be summarized as follows. Although the concentrations of alternative biocides in sediment samples in this study were generally lower than those in other aquatic areas, the concentrations of these compounds in sediment were higher than those in the other aquatic areas. Furthermore in regard to the horizontal distribution, although there was no major difference in the concentrations of alternative biocides concentrations in water samples in this study, the concentrations of most alternative biocides in the sediment samples from Sts. H1-H2 and O1 which were in proximity to the fishing and industrial areas, were higher than those at the other stations (Table 5-6). This trend indicates that the use of alternative biocides is currently decreasing, although alternative biocides were widely used on ship hulls in the past. In addition, the concentrations of most alternative biocides were lower than those before TBT ban by IMO, suggesting that the alternative biocides which measured in this study are no longer in use, and their compounds may have been replaced the other alternative compounds such as pyrithions and borans.
Partition coefficient between water and sediment samples
The partition coefficients (Ksw) between the water and sediment of each alternative biocide were calculated by dividing the concentration of each alternative biocide in the sediment by the concentration of each alternative biocide in the bottom water (Fig.5). Some substances not detected in water samples are derived from this calculation. The Ksw of diuron, Sea-Nine 211, Irgarol 1051, M1, dichlofluanid and chlorothalonil were 13 x 104, 21 x 104 - 53 x 104, 2.5 x 104 - 12 x 104, 1.7 x 104 - 7.9 x 104, 0.12 x 104 – 27 x 104 and 17 x 104 – 14 X 106, respectively. Harino et al. (2005) reported Ksw before the ban of TBT. The ratios of concentrations for Diuron, Sea-Nine 211, Irgarol 1051, and M1 were 0.27 x 104, 0.069 x 104, 0.03 x 104, and 0.087 x 104, respectively. The Ksw of Diuron, Sea-Nine 211, Irgarol 1051, and M1 in this study were higher than those before ban of BTs, although concentrations of these compounds in water samples were lower than those before the ban on TBT, suggesting that the degradation rate of these compounds accumulated in the sediment is slow. The Ksw of alternative biocides were similar to those of BTs.