International Journal of Scientific Research in
Environmental Science and Toxicology

Chlorine is extensively used as a powerful oxidizing agent in the countries surrounding the Arabian Gulf for water treatment and biofouling control. Its usage has been increasing significantly as demand for water grows considerably both in industry and for domestic use. This is due to the fact that it is a well-tested technology, has had a history of long-term worldwide industrial use and is of acceptable cost. While the Arabian Gulf waters support a range of coastal and pelagic marine habitats including mangrove forests, seagrass meadows and coral reefs, marine organisms in these waters are living close to their tolerance limits due to the extreme environmental stressors like temperature and salinity. Anthropogenic stressors such as chlorine may further exacerbate these natural stressors. In seawater, chlorine produces a mixture of hypochlorous acid and hypochlorite ion. These rapidly react with the bromide ion to form a mixture of hypobromous and hypobromite ion. Total residual oxidants formed by chlorination although are short lived and not persistent in seawater, they can be quite toxic. In the present study, toxicity data were obtained from 7 acute toxicity tests and 3 chronic toxicity tests using Arabian Gulf aquatic species from different trophic levels. The study also examined the effect of temperature and developmental stages on toxicity of chlorine. Furthermore, differences in the species sensitivity distribution between native and non-native species were compared. The main finding of the study showed that there was no significant difference between native and non-native species for chlorine toxicity. This would suggest that toxicity data from different geographic region can be used in deriving site- specific ecological risk assessment of chlorine.

Keywords: Chlorine; Arabian Gulf; acute toxicity; chronic toxicity; risk assessment; species sensitivity distribution;

The Arabian Gulf is a relatively small and isolated body of water [Figure 1], and the activities occurring in one country are likely to impact the state of marine ecosystems throughout the region [57] [4]. Gulf waters support a range of coastal and open-water marine habitats, and high levels of salinity and surface temperature are the main natural stressors. In addition, the Arabian Gulf has high levels of anthropogenic activity [28] [15] Such activities include reclamation, industrial and domestic sewage effluent discharge, as well as hypersaline water discharge from desalination plants [49] [12] [63]. The region has also experienced significant oil pollution as a consequence of regional conflicts [4] [58]. The Gulf’s vulnerability to environmental stressors led to a series of regional actions in the 1980s [3] [16] and despite recent improvements, environmental regulation and management have typically lagged behind the rapid pace of development across the region. Consequently, it is important for decision-makers to consider sustainability as a necessary facet of economic development to realize improved management that can support conservation and recovery of these important ecosystems. To this end, scientific data and information on both the ecosystems and activities (human and natural) taking place within the region are critically important. Additionally, establishing a Gulf-based transnational regulatory and enforcement framework backed by these data and applied research could be very beneficial.

Discharges of total residual oxidant (TRO), referring to the sum of the free chlorine / bromine and combined chlorine /combined bromine, from the disinfection and anti-fouling processes are common due to the continued application of chlorine in many municipal water treatment and industrial facilities. TRO has the potential to be toxic to aquatic organisms and thus should be examined to understand potential impacts from discharge. Numerical water quality guidelines (WQG) for the protection of aquatic life for chlorine have been published by the U.S. EPA [30] and Environment Canada [18] and draft guidelines are available from the European Commission [29]. While these guidelines consider available toxicity data for TRO and in some cases the same data sets, each of them produced different numerical values. The U.S. EPA reported TRO guidelines of 11.5 and 7.55 μg/L for freshwater and saltwater species, respectively. The CCME reported a WQG for reactive chlorine species of 0.5 μg/L for freshwater and a WQG for chlorine-produced oxidants of 0.5 μg/L for marine water. The European Commission reported a TRO probable no-effect concentration (PNEC) of 0.06 μg/L and a free available chlorine (FAC) PNEC of 0.04 μg/L for both freshwater and saltwater organisms.

Guidelines established by the Qatari Ministry of Environment (MoE) (one of the Arabian Gulf states) for TRO and CBPs in cooling discharge waters specify that the maximum concentration of free TRO is 0.05 mg/L at the discharge point [51]. However, greater than 0.05 mg/L is permitted if a site-specific dispersion model is developed to demonstrate that the chlorine concentration does not exceed 0.01 mg/L at the edge of the mixing zone. This value is in line with the USEPA TRO guideline for marine species [1].

When elemental chlorine is added to water (e.g., wastewater, freshwater), it hydrolyzes into hypochlorous acid (HOCl) and hydrochloric acid (HCl). Hypochlorous acid is a weak acid and dissociates into hypochlorite ion (OCl-). Similarly, when calcium or sodium hypochlorite is added to the water, the salts dissociate to form hypochlorous ions. The relative proportion of the three chlorine species, elemental chlorine, hypochlorous acid and hypochlorite ion, depends on pH and to a lesser extent on temperature [52]. The chemistry of chlorine exposure to seawater is complex [66] [21] [55]. In seawater, chlorine produces a mixture of hypochlorous acid (HOCl) and hypochlorite ion (OCl−). These rapidly react with the bromide ion to form a mixture of hypobromous acid (HOBr) and hypobromite ion (OBr−). The acute chlorinated oxidants are short lived and are not persistent in seawater. The environmental concern of chlorination is the production of numerous and more persistent, compounds formed by complex reactions between chlorine/bromine and the organic constituents of seawater, collectively described as chlorination by-products (CBPs) [2]. Many CBPs are persistent and may be toxic to marine organisms subjected to long-term exposures [8]. The terms chlorine refers to the sum of the free chlorine / bromine and combined chlorine /combined bromine. For our purposes, we will refer chlorine to combined oxidants as defined above.

Objectives for this work were to [1] run acute and chronic toxicity tests of chlorine on marine species native to Qatar and the Arabian Gulf, [2] correlate sensitivity of these species to nonnative marine organisms used as indicator species in WQG, [3] extrapolate research findings from this study to explore ways to inform and enhance environmental management activities in the Arabian Gulf coastal areas.

Analytical grade calcium hypochlorite (with 75% available chlorine) was purchased from BBH Prolabo, UK. Artificial seasalt was obtained from Seachem, US and water was prepared by reverse osmosis. All other reagents used in this study were of high purity. Chlorine test solution was prepared by diluting calcium hypochlorite in filtered seawater. Aliquots of the treatment solution were sampled periodically to monitor the residual chlorine concentration. The concentration of chlorine was determined by N, N-diethyl-p-phenylenediamine (DPD) method [40], using a Hach DR2800 (Hach, USA) [36]. Fluorescein diacetate was used for algal cell viability [26] and was purchased from Sigma (St. Louis,Mo, USA). Bloodworms and brine shrimps (Artemia sp.) used for feeding killifish were purchased from Aqua Art, Doha, Qatar.

In our study seven local aquatic species from the Arabian Gulf were chosen for the acute and chronic toxicity tests. See Table 1 for testing conditions.

Two microalgae species, Synechococcus Sp. and Chaetoceros Sp. were isolated from the coastal waters of north west of Qatar and their cultures were established and maintained in the laboratory using F2 medium [35] and filtered seawater adjusted to 40 ppt. Culture flasks containing100 ml of autoclaved culture medium were inoculated with about a 1 ml of stock culture and placed at 28 °C under simulated natural light conditions (12 h light: 12 h dark photocycle). The viability of algal cells after treatment was analyzed by flow cytometry with FDA staining [32]. Esterase activity was recorded by detection of emitted fluorescence signals with the application of a blue laser light. FDA, which is taken up by live cells and converted to its fluorescent derivative Fluorescein by cellular esterase, was applied after dilution in acetone (5 mg per mL). Staining time was 5 min at room temperature in darkness. The esterase activity of the cells was evaluated using a BD Accuri C6 flow cytometry system (BD Biosciences, USA) with a 488 nm excitation argon laser equipped with fluorescent emission filters (FL1-525 nm). The maximum FDA fluorescence emission was detected at 525 nm.

Toxicity tests with the crustaceans Euterpina acutifrons and Microstella Sp., were performed on nauplia and copepodite stages. Methods are described in detail in [50]. Mature females were isolated from coastal waters of North West of Qatar and pure cultures were established in the laboratory. For the toxicity tests, cultures were synchronized to procure same age group individuals. Test organisms were fed with a mixture of algae for an hour prior to the test. Five to seven copepodites were picked with a flame-rounded pipette guided with a stereo microscope and placed in test vials containing the test solutions.

General methods for invertebrates followed [10] [27]. Adult Pearl oysters (Pinctada radiata) and sea urchins (Echinometra mathaei) were collected from North West Qatar and acclimatized to laboratory. Oysters and Urchins were fed with a mixture of laboratory grown microalgae and commercially obtained seaweeds, respectively. Spawning was induced by dissection for the pearl oyster and by injecting potassium chloride for the sea urchins. Gametes collected from the spawners were suspended in 40 ppt filtered sea water and were checked under the microscope for maturity. Male and female gamete suspension were mixed in a beaker and stirred gently to allow fertilization for oysters.

For sea urchins the eggs were collected and counted under the microscope and a suspension of 2000 eggs per ml was made. 1 ml of the egg suspension was transferred to 3 ml well plates. This was followed by an addition of sperm suspension. Simultaneously the toxicant was added to the well plates to check the impact of the toxicant on fertilization. Four replicates per treatment were used. After 8 h incubation at 23 °C in an isothermal chamber, contents in the wells were preserved using formalin for later observations under an inverted microscope. Successful fertilization in sea urchin is indicated by the formation of a fertilization membrane around the egg. The recorded endpoint for pearl oyster was the percentage of normal D-larvae in samples of a minimum of 100 individuals per well. A larva was considered developing normally when the shell was D-shaped (straight hinge) and the mantle did not protrude out of the shell. Fertilization success and normal larvae development in the control, for both species, were always more than 70%.

Killifish embryo, juveniles and adult bioassays were performed following OECD guidelines No 236 [47], 215 [46] and 203 [45]. Fifty fertilized eggs (blastula stage), 10 individual juveniles and 10 individual adults were carefully distributed to the exposure chambers. Test chambers were incubated in a biological incubator, see [Table 1]. The observations were made using a thick slide with a concave chamber, which was filled with clean seawater. Each embryo was carefully placed in the chamber and observed under a binocular dissection microscope, embryos were considered dead when no heartbeat could be observed. Survival after 96 h was the recorded response in juvenile and adult acute tests and for embryo hatchability and survival after 10 days. Mortality in the controls was always below 10%.

All tests were static renewal, unless otherwise stated, whereby test solutions were spiked with chlorine at 24 h intervals. Measured chemical parameters of dilution water were as follows: pH 7.8 ± 0.4, dissolved oxygen (DO) 6.8 ± 0.5 mg/L, total organic carbon 0.02 mg/L, and salinity 40 ± 1 ppt. The toxicity tests were conducted in triplicates, at assigned concentrations, and blank control. All tests were undertaken at a light: dark photoperiod of 12: 12 h. Test organisms were not fed during the acute test periods. Test chambers were kept inside an incubator adjusted to maintain the water temperature at 23 ± 2 °C, unless otherwise noted. Temperature, DO, and pH were measured in test chambers daily in the acute toxicity tests and at least once a week in the chronic toxicity tests. Biological observations were performed at least once daily for all studies.

In the acute toxicity test, 72-h-EC50 (effective concentration in 50% of the test organisms over 72 h) for microalgae, 48-h-LC50 for copepod, pearl oyster and sea urchins and 96-h-LC50 (lethal concentration in 50% of the test organisms over 96 h) for fish larvae and adults and 10-d- LC50 for fish embryos were used as main endpoints. In the chronic toxicity test, 21-d-EC10 (effective concentration in 10% of the test organisms over 21 days) for copepod and 120-d-EC10 (effective concentration in 10% of the test organisms over 120 days) for oysters and 28-d-EC10 for fish.

Prior to the toxicity tests, all test organisms were acclimated to general test conditions for a minimum of 7 days. Acute toxicity tests set up of chlorine to seven local aquatic organisms are shown in [Table 2, 3].

The fish used in the chronic toxicity test were the same as in the acute toxicity test, i.e., chronic study was a continuation of the acute study. Twenty eight days chronic toxicity tests were conducted. The fish were fed with bloodworms at a rate of 0.1% body weight twice daily. Endpoints observed included survival, and change in total length and wet body weight. Nominal exposure concentrations in the test were 0.00 (control), 0.025, and 0.1, mg/L chlorine, delivered through a pump to dose chlorine at regular intervals so as to maintain the test concentrations and to account for loss of chlorine due to volatilization.

Twenty-one-day life table tests using nauplii of Euterpina acutifrons (< 24 h age) were conducted in 1000 mL beakers filled with 300 mL test solutions. The copepods were fed once every day with a mixture of green algae and diatom with a total concentration of 1.0 × 105 cells/mL in the test solution. The survival, moulting, growth and reproduction were recorded every alternate day. The endpoints include the time to first brood, number of broods and number of exuviations. Nominal concentrations in the definitive test were 0.00 (control), 0.025, and 0.1mg/L chlorine, delivered through a pump to dose chlorine at regular intervals. The toxicity tests were carried out in two replicates with each starting with 30 first stage nauplii.

The oysters used in the chronic toxicity test were the same as in the acute toxicity test. 70 healthy oysters were collected and distributed into tanks that were fully aerated and filtered through a biological filtration system. One hundred and twentyday chronic toxicity tests were conducted. The oysters were fed with a mixture of algae daily. Endpoints observed included survival, growth (length in cm) and body weight (g) gain. Nominal exposure concentrations in the definitive test were 0.00, 0.025, and 0.1, mg/L chlorine, delivered through a pump to dose chlorine at regular intervals.

In this experiment, Synechococcus sp., Euterpina acutifrons, and embryos of Aphanius dispar were subjected to different temperatures (16, 23, 27, 32 and 38 °C) to study the impact of temperature on toxicity of residual chlorine. Conditions used were similar to that of acute and chronic toxicity studies.

More than 125 references were reviewed in order to compare with the results of the present study. The datasets considered for comparison were those which had test conditions, recorded chlorine measurements, and used either semi-static renewal or a flow-through system of dosing chlorine. In addition to the test method, other criteria for selecting existing data sets include measuring standard parameters of dissolved oxygen, temperature, pH and salinity and achieving acceptable criteria in the control.

Probit methodology was employed to calculate the 48-h- EC50, 96-h-LC50, 10-d-LC50, EC10 for chronic endpoints values and corresponding 95% confidence intervals. JMP- SAS 13 (SAS Campus Drive, Cary, NY) and GraphPad Prism 7 (San Diego, CA) were used for data analysis and display.

In the present study, toxicity data were obtained from 7 acute toxicity tests and 3 chronic toxicity tests using seven Arabian Gulf native species from different taxonomic levels as shown in [Table 4] Mortality in controls were within published standard guidelines for each species. Results of acute toxicity tests showed that copepodites of Euterpina acutifrons with a 48-h-LC50 of 0.076 mg/L and embryos of Pinctada radiata with a 48-h-LC50 of 0.073 mg/L were the most sensitive species to chlorine followed by Synechococcus Sp., Chaetoceros Sp., Echinometra mathaei, Euterpina acutifrons nauplii, Aphanius dispar juveniles and Aphanius dispar adults. The least sensitive species were found to be the copepodites of Microstella Sp. with a 48-h-LC50 of 1.004 mg/L and Aphanius dispar embryos with 10-d-LC50 of 1.017 mg/L.

In this experiment, semi-static versus flow-through dosing methods were examined on Aphanius dispar embryos, larvae and adults. The difference in toxicity of the flow-through was at least 3 to 6 times more in magnitude than the semi-static [Table 4] for embryos the LC50 went up from 0.074 to 0.454 mg/L and for juveniles 0.059 to 0.215 mg/L. In addition, LC50 differed depending on the developmental stages of the species that was exposed to chlorine. In the case of Euterpina acutifrons, the nauplii were less sensitive than the copepodite, 0.217 and 0.076 respectively. The fish juveniles were found to be more sensitive than adults, 0.217 and 0.743 respectively. In most cases embryos were less sensitive than larvae or adult stages and this is mainly attributed the protection provided by the outer membrane, the chorion.

The impact of temperature on chlorine toxicity was tested on three native species [Figure 2] and was later compared to data obtained from non-native species [Figure 3] At either low (16°C) or very high temperatures (>38°C), chlorine toxicity was not tested due to impact of these temperatures on the well-being of test species. However, on temperatures used 23, 27, 32 and 38°C there was no significant difference observed in chlorine toxicity with the concentrations used. Nevertheless, depending on species, at intermediate temperatures (~ 23-27°C) chlorine showed an elevated toxic effect on all test species.

Chronic toxicity values of chlorine to 3 aquatic species are shown in [Table 5] Results obtained in this study and those of previous studies using non-native species (Table 6) were compared. In the present study we found that the lowest observed effect concentration (LOEC) for 21d for planktonic crustacean Euterpina acutifrons to be 0.026 mg/L. The 28-d-EC10 for growth of fish Aphanius dispar was found to be 0.020 mg/L. In general, the sensitivities of the native species tested in this study were similar to those reported in previous studies of non-native species.

For the present study, the species sensitivity distribution (SSD) based on both native and non-native species was generated. Since there has been limited information on the toxicity of chlorine to native and non-native species, comparison could only be conducted on the SSDs constructed from acute toxicity data of native and non-native [Table 4, 7] species. Based on the comparison in this study, the SSD curves [Figure. 4] indicated that the native Arabian Gulf organisms with acute exposures to chlorine were as sensitive as the non-native species that have been used in deriving WQC elsewhere. The HC5s were 0.054, 0.039, and 0.045 mg/L, respectively.

The purpose of the present study was to conduct acute and chronic chlorine toxicity tests on native species of the Arabian Gulf region and to compare sensitivity to non-native species used WQC. Based on the present study’s results, the EC50 and LC50 observed for the native species were closely similar to those reported for non-native species with similar test conditions. All the studies (including this study) on chlorine acute toxicity shows that toxicity of chlorine depends on exposure methods, environmental conditions and stage of development of the test species. The lowest LC50 reported by Cairns 1978 was 0.076 mg/L (chlorine) at 25°C. Similarly, Thatcher [61] has investigated acute toxicity to seven invertebrates of which two shrimps (Pandalus goniurus and Crangon niricauda) were found to be most sensitive to chlorine. The 96h LC50 of Pandalus goniurus was 0.09 mg/L.

Furthermore, [37] reported a 96-hour LC50 value of 0.06 mg/L to rainbow trout (Salma gaidneri), which contrasts with LC50 values of 0.15 mg/L for trout (Salvalinus fontinalis) and 0.032 mg/L for salmon (Oncorhynchus kisutch), reported by [61]. The effects of chlorine to 11 phytoplanktonic algae species were tested by [33] and report the chlorine concentrations causing 50% growth reduction in a series of 24 h static tests; the LC50values ranged from 75 to 330 μg/l. The acute toxicity of chlorine in fish has been the most studied, with the most referenced including (61] [37] [65]. The lowest 96h LC50 value was determined by [61] for juvenile Oncorhynchus kisutch (Coho salmon, LC50= 0.032 mg/L) and juvenile Oncorhynchus gorbuscha (pink salmon, LC50 = 0.023–0.052 mg/L). Exposing fish to chlorine adversely impacts the structure of the gill with an increase in mucus production around the gill and destruction of the respiratory wall epithelium [11] [19] [65]. Exposure to sub-lethal levels of chlorine to bluegill and rainbow trout caused histological damage to the gills ultimately resulting in death by asphyxia [11].

The results of [59] demonstrated the effect of exposure methods i.e. flow through versus static systems. The LC50 for the crustacean, Ceriodaphnia dubia in a 24 hour flow through test at 25°C was equal or greater than 0.005 mg/L. The LC50 determined under static conditions was 0.048 mg/L which is in the range of the values reported in 17 at the same temperature. Essentially a 10-fold increase in LC50 was observed, driven mostly by the difference in exposure condition. However, semistatic exposure was the method of choice in this study as this is more representative of how industries operating within the Arabian Gulf administer chlorine to counter act the biofouling in the cooling pipes.

Studies have reported that early developmental stages of fish and invertebrates were more sensitive to toxicants than the adults [38] [40] [54] [42]. The variability in sensitivity may be due to several factors; surface area/volume ratio (particularly with juvenile fish); greater uptake of toxicant from environment; under developed homeostatic mechanism to deal with toxicants; immature immune system and under developed organs (liver and kidney) which has an important role in detoxification and elimination of toxicants. For fish, it has been reported in this study and in others [62], that the juveniles are more sensitive than embryos. The difference in the effect of chlorine between embryos, juveniles and adults reported in this study could be attributed to the chorion, the membrane surrounding the egg. As previously demonstrated the chorion is believed to provide a barrier against substances in early stages of embryo development because of the water hardening of the chorion, allowing the egg to become mechanically more resistant and possibly restricting the entrance of waterborne substances into the perivitelline fluid.

Temperature affects the physiology of organisms and there is a point when the temperature is too low or too high that death may occur. Organisms have a range of temperature that they can tolerate without any adverse effects. When the temperature changes in the environment, all the cold-blooded organisms would adjust their body temperatures to be equal to the external temperature. The rate of heat exchange in these organisms is very rapid and as such their metabolic rate will increase. 23 showed that metabolic rates increase with an increase in temperature for tallest fish. The relationship between metabolic rate and temperature is often expressed as Q10, which measures the rate of increase for every 10ºC rise in temperature. For coldblooded animals, the Q10 ranges from a factor of 2 to [23]. Since temperature alone can be lethal, when an organism is exposed to a chemical at an elevated temperature, a synergistic effect could occur and enhanced toxicity may be observed.

However, there was a common observation among species that between 23-27°C there was a noticeable relative increase in toxicity. This temperature is more than the average measured (22°C ± 0.2) in the Arabian Gulf. At high temperatures such as 27°C the metabolic rate of the species would be expected to increase, which could have enhanced the toxicity as found in this study showing a synergistic effect of both chlorine and temperature. Furthermore, at higher temperatures (higher than 32°C) increased evaporation of available chlorine in the waters might have reduced exposure in the animals. At low temperatures, 16°C, the metabolic rate of the animal decreases to a lower level which would lower the negative impact of chlorine.

In general, the sensitivities of the native species tested in this study were similar to those reported in previous studies of nonnative species. It is in range with previous studies which reported that toxicity value for planktonic crustacean Ceriodaphnia dubia was 0. 056 mg/L [59] for 10d lifecycle test. [9] reported that 21-d no observed effect concentration (NOEC) for the survival of Pimephales promelas was 0.026 mg/L. The 28-d-EC10 for growth of fish Aphanius dispar in this study was 0.026 mg/L, and previous studies reported that 28 days NOEC for survival of Menidia peninsulae and Oryzias javanicusb were 0.04 and 0.078 mg/L, respectively [34].

Based on the comparison in this study, the SSD curves [Figure. 4], indicated that the native Arabian Gulf organisms with acute exposures to chlorine were as sensitive as the non-native species that have been used in deriving WQC elsewhere. The HC5s were 0.054, 0.039, and 0.045 mg/L, respectively. Previous study found that natural history, habitat type and geographical distribution of the species used to construct the SSD did not have a significant influence on the assessment of hazard, and this was in accordance with this study [39] [31).

This study is a contribution in the assessment of the effect of chlorine in the aquatic environment. Toxicity values of 7 acute and 3 chronic tests for seven native species were obtained in this study, among which planktonic crustacean, Euterpina acutifrons was found to be the most sensitive species. Comparing the species sensitivity distributions, there was no significant difference between native and non-native species. This indicates that toxicity data from different geographic region may have validity in site-specific ecological risk assessment of chlorine.

For the present study, the species sensitivity distribution (SSD) based on both native and non-native species was generated. The question was brought up [26] [67] about the feasibility of using toxicity data of species from one geographical region to assess the ecological risk posed to species in a different region. Moreover, differences in the sensitivity of cold-water, temperate, and tropical fish species have been reported [6] [25] [67]. Since there has been limited information on the toxicity of chlorine to native and non-native species, comparison could only be conducted on the SSDs constructed from acute toxicity data of native and non-native [5] [43] [64].

This work has been funded by ExxonMobil Research Qatar. We thank the Environmental Management Team for their support during toxicity tests.

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