Effect of brine discharge from a desalination plant on macrobenthic communities in the NW Mediterranean

MARINE ENVIRONMENTAL RESEARCH Marine Environmental Research 62 (2006) 1–14 www.elsevier.com/locate/marenvrev

Effect of brine discharge from a desalination plant on macrobenthic communities in the NW Mediterranean N. Raventos *, E. Macpherson, A. Garcı´a-Rubie´s Centro de Estudios Avanzados de Blanes (CSIC), Carr. Acc. Cala Sant Francesc 14, 17300 Blanes (Girona), Spain Accepted 15 February 2006

Abstract This paper examines the possible effects of discharges from a desalination plant on the macrobenthic community inhabiting the sandy substratum off the coast of Blanes in Spain (NW Mediterranean) using multivariate and univariate analyses. Two controls and one putatively impacted location were selected and visual censuses were carried out 12 times before and 12 times after the plant had begun operating. No significant variations attributable to the brine discharges from the desalination plant were found. The failure to record any impact may be explained by the high natural variability that is a characteristic feature of bottoms of this type and also by the rapid dilution undergone by the hypersaline brine upon leaving the discharge pipe.  2006 Elsevier Ltd. All rights reserved. Keywords: Desalination; Environmental impact; Macrobenthic community; Mediterranean Sea

1. Introduction The number of desalination plants in regions with high water deficits has been growing of late, particularly in the more southern countries of the northern hemisphere, which have seen their needs for fresh water for their population, industry, and agriculture climb continually (Allam et al., 2003; Al-Agha and Mortaja, 2005). Whether the process used is *

Corresponding author. Fax: +34 972337806. E-mail address: [email protected] (N. Raventos).

0141-1136/$ - see front matter  2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.marenvres.2006.02.002

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N. Raventos et al. / Marine Environmental Research 62 (2006) 1–14

reverse osmosis or distillation, desalination plants generate volumes of hypersaline brine which are then discharged back into the sea and which may, therefore, affect littoral marine communities (Ho¨pner and Windelberg, 1996). The repercussions of high salinity levels on marine ecosystems and organisms can take a variety of forms (e.g., Young and Potter, 2002; Moscatello and Belmonte, 2004; VegaCendejas and Hernandez-Santillana, 2004). Nevertheless, despite the growing number of desalination plants, there have been very few studies dealing with the impact of hypersaline brine inputs on benthic communities. The few reports available indicate that brine discharges have led to reductions in fish populations and to plankton and coral die-offs in the Red Sea (Mabrook, 1994), to mangrove and marine angiosperm mortalities in the Ras Hanjurah lagoon in the United Arab Emirates (Vries et al., 1997), as well as to high levels of copper and nickel pollution in the sediment off Key West, Florida (Chesher, 1975). In addition, Posidonia oceanica has been observed to have very low tolerance to increased salinity in the Western Mediterranean (Buceta et al., 2003). Considering the possible repercussions ensuing from any type of discharge, e.g., sewage, brine, thermal discharges, requires a suitable sampling design and selection of the communities that may be affected (Guidetti and Bussotti, 2002). Naturally occurring spatial and temporal variations can make it extremely difficult to detect human impacts (Roberts et al., 1998), hence a variety of sampling designs have been developed for the purpose of detecting and assessing the effects of anthropogenic inputs. Green’s (1979) BeforeAfter-Control-Impact (BACI) design proposed collecting samples before and after a planned impact in order to be able to assess differences between the disturbed area and a control site over time. This design, however, may confuse the effects of the impact with other types of unique natural fluctuations that occur at one site but not at another (Hurlbert, 1984; Stewart-Oaten et al., 1986). More recent designs have recommended sampling at several sites on a number of occasions before and after the impact to achieve suitable spatial and temporal replication (Underwood, 1992, 1994). In the present study some spatial and temporal descriptors of the communities of certain organisms were evaluated before and after start-up of the desalination plant in order to achieve the necessary spatial and temporal replication (Underwood, 1992, 1994). The macrobenthic community consisting, for example, of fishes, decapod crustaceans, molluscs, echinoderms, etc., was selected because it could be readily sampled by means of underwater visual methods (Harmelin-Vivien et al., 1985; Macpherson and Raventos, 2004). The aim of our study was, therefore, to examine the possible effects of discharges from a desalination plant on the macrobenthos inhabiting the sandy substratum off the coast of Blanes in Spain (NW Mediterranean).

2. Materials and methods 2.1. Study area The discharge pipe was located off the mouth of a small (4 Hm3/year; 1 Hm3 = 106m3) temporary river in the NW Mediterranean Sea (4139 0 N, 247 0 E). The volume of the river run-off may vary by orders of magnitude both within a given year and between years. Flows tend to be low in the summer, when the river bed is usually dry for several months at a time (Merseburger et al., 2005).

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The coastline is open, and the 30-m isobath is located barely 500 m from the shore, at which point the bottom drops off steeply. The head of a submarine canyon approaches the coast and allows oceanic water to penetrate inshore. The substratum consists mainly of coarse sandy sediments 500–750 lm in size with some medium sand (250–500 lm) mixed in (Pinedo et al., 2000). Sediment composition is similar both north and south of the river mouth. The benthic fauna inhabiting the area has been characterized in detail by a number of workers (Sarda` et al., 1999, 2000; Pinedo et al., 2000; Macpherson and Raventos, 2004). The sampling site located within the potentially impacted area near the river mouth has been designated D, for discharge pipe. It comprises the narrow area along the pipe where discharge outlet effects have been detected (=10 m). Two control sites located north and south of the pipe (designated C1 and C2, respectively) were chosen because they were topographically similar to but situated a considerable distance (>1000 m) away from the supposedly impacted location and thus were not exposed to the effects of the brine discharges. 2.2. Characteristics of the brine discharge The brine discharge pipe was L-shaped. An initial section of pipe was buried and ran perpendicular to the coast. The pipe emerged at a depth of 16.5 m and continued on for another 74 m, where it was connected to a second section of pipe that ran for 95 m nearly parallel to the coast. The junction between the two pipe sections was located at a depth of 19 m, the deepest depth attained. The outlet portion of the pipe (169 m in total), from where it emerged from the sediment, was a diffuser that had 43 perforations along its length so that the discharges would be diluted quickly. The volume of seawater collected for desalination is 22 Hm3/year, yielding 10 Hm3/year of potable water. The salt concentration values in the brine produced by the plant are around 60 g/l, and the estimated brine discharge is 12 Hm3/year. Salt concentrations at the discharge outlet fall off sharply with distance from the pipe, with salinity values being back to normal for the area at a distance of 10 m from the outlet pipe. The pattern for nutrient (NH4, NO3, SiO4, PO4) concentrations was similar to that for salinity. 2.3. Sampling procedures The transect technique was chosen as the most appropriate visual census method for both smaller specimens and fast-swimming species (Harmelin-Vivien and Francour, 1992; Garcı´a-Rubie´s and Zabala, 1990). Visual censuses were effected monthly from February 2002 to January 2003 before start-up of the desalination plant and from May 2003 to April 2004 after the plant had begun operating. Sixteen visual transects measuring 2 · 50 m each were monitored by two scuba divers at each site (D, C1, and C2) monthly. Transects at D were performed always along the pipe in order to make sure of being in the discharge affectation area. Specimens of each species sighted (in number) were recorded on plastic tablets according to visual monitoring techniques commonly used in fish studies (e.g., Harmelin-Vivien et al., 1985; Macpherson et al., 2002). The taxonomic categories considered included Pisces, Echinodermata, Mollusca, Polychaeta, and Decapoda Crustacea. The latter were mainly hermit crabs of the genera Anapagurus and Pagurus, which ordinarily spend most of their time buried in the sand, and a small hand net was used to trawl the bottom along the transects to collect and count all the specimens present

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(Macpherson and Raventos, 2004). During settlement periods recruits of fish species were also censused specifically. 2.4. Data analysis Quantitative spatial and temporal differences in the macroepifaunal counts obtained from the visual censuses were plotted by sampling site and time (before/after) using non-metric multidimensional scaling (MDS) on a two-dimensional ordination diagram. The values were transformed to sqrt(x) so that each species would contribute more evenly to the analyses (Clarke, 1993). Differences between sampling times and sites were assessed using an analysis of similarities (Jaccard’s index of similarity; two-way crossed ANOSIM). A one-way analysis of similarity was used to compare each site before and after plant start-up. The SIMPER program was then employed to identify the contribution of each individual taxon to the global dissimilarity (part of the PRIMER package from Plymouth Marine Laboratory, UK). This procedure has previously been used to assess local differences after disturbances to various taxa (e.g., fishes by Guidetti and Bussotti, 2002). The descriptors employed for the community were species richness, species abundance, and the Shannon–Wiener diversity index. Differences between the three descriptors and between the mean abundance values for selected species (Pagurus excavatus, Xyrichthys novacula, and Trachinus araneus) before and after desalination plant start-up were assessed by a beyond BACI design involving asymmetrical analyses of variance. This technique has been described in full by Underwood (1993) and Glasby (1997). The samples were collected according to two time scales, namely, before and after plant start-up (B) (Fixed factor) and at times within each before or after period, T(B) (random, nested factor). Cochran’s test was used to test for homogeneity of variances. Values were transformed to ln(x + 1) or sqrt(x) in a few cases, when necessary. 3. Results The divers recorded up to 29 species belonging to nine different taxa (Table 1). Total species richness varied slightly among the sites, being slightly higher at C1 (29 species in 2002 and 27 species in 2003) than at C2 (23 species in 2002 and 23 species in 2003) and D (27 species in 2002 and 21 species in 2003). Qualitative affinities between the sites were generally high, in all cases above 70%. The greatest before-after differences occurred at sites C1 (a similarity index of 75.0) and D (a similarity index of 79.2), the lowest difference being recorded at site C2 (a similarity index of 82.6). The MDS diagram for centroids (Fig. 1) was indicative of some difficulty in displaying the relationships between the samples (moderately high stress at 0.23). Maximal variation was due to changes in Pagurus excavatus, Anapagurus petiti, and Anapagurus alboranensii abundances. Also, differing abundance levels for Pomatoschistus marmoratus and, to a lesser extent, Xyrichtis novacula, exhibited pronounced differences due to the uneven recruitment of these two species. Comparison of the before and after variations at each site revealed the smallest differences to have taken place at C1 (R = 0.138; p = 0.02) and the largest to have taken place at D (R = 0.366; p = 0.002) but also at C2 (R = 0.324; p = 0.001).

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Table 1 Species list sampled by visual censuses Mollusca Gastropoda Aplisia spp Bolinus brandaris Cephalopoda Sepia officinalis Octopus vulgaris Annelida Polychaeta Spirographis spallanzanii Mesochaetopterus rogeri Arthropoda Crustacea Dio´genes pugilator Pagurus excavatus Anapagurus petiti Anapagurus alboranensii Thia scutellata Parthenope massena Echinodermata Asteroidea Astropecten aranciacus Astropecten jonstoni Ophidiaster attenuata Phiuroidea Ophioderma longicaudum Echinoidea Sphaerochinus granularis Echinorcardium cordatum Spatangus purpureus Holothuroidea Holothuria spp Cerianthus membranaceus

Chordata Osteichthyes Arnoglossus kessleri Bothus podas Echelus myrus Gobius paganellus Hippocampus hippocampus Ophisurus serpens Pagellus erythrinus Pomatoschistus marmoratus Serranus cabrilla Serranus hepatus Synodus saurus Trachinus araneus Trigla lucerna Uranoscopus scaber Xyrichtis novacula

Fig. 1. Non-metric multidimenzsional scaling (MDS) ordinations of centroids comparing macrobenthic communities from a putatively impacted location (D = s) and control (C1 = h and C2 = n) locations. Before = open symbols, After = filled symbols (2002 and 2003 respectively).

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Analysis of the percentage similarities before and after desalination plant start-up showed the variations to be due mainly to decreases in Pomatoschistus marmoratus, Pagurus excavatus, and Echinocardium cordatum and to slight increases in Anapagurs alboranensii and Mesochaetopterus rogeri. The variations in the sqrt transformed abundance values for these species alone explained 57.6% of the before and after dissimilarities. This same pattern was observed at all the sites, including a pronounced decrease in Pomatoschistus marmoratus at C2 and a sharp drop in Pagurus excavatus at D, where this species virtually disappeared, after the brine discharges commenced. Even though these before and after disparities seem appreciable, the analysis of variance results indicated that in no case could the differences be attributed to impacts from desalination plant operation. Accordingly, abundance values for Pagurus excavatus and Anapagurus alboranensii and Anapagurs petiti (Table 2) underwent substantial variation before and after the brine discharges started at both the control sites as well as at the presumably impacted site (Fig. 2). This means that the differences between the control sites and the impacted site disclosed by the comparison were not significant. The analysis of variance for the mean abundance of the polychaete Mesochaetopterus rogeri and the fish Trachinus araneus. (Table 2 and Fig. 2) yielded very similar results. Indeed, the only real variation (p = 0.176) found between the supposedly impacted site and the control sites after the disturbance was a certain difference in the case of Xyrichthys novacula (Table 2 and Fig. 2). Even though the difference was not significant, it nonetheless gave rise to appreciable variation (p = 0.076; see Table 2) before and after onset of the brine discharges at the supposedly affected site (Fig. 2). These findings indicate that there was no significant short-term impact in the mean abundance values for these species, principally as a consequence of the high levels of natural variability recorded at the control sites. No significant differences were observed for recruit abundance (Xyrichtis novacula, Bothus podas, Pomatoschistus marmoratus) between the sites. Similarly, no significant effects were observable among the community descriptors (species richness, diversity index, species total abundance) [Table 3 and Fig. 3]. 4. Discussion The findings compiled in this study suggest that the communities considered are subject to high temporal and spatial variability, which is typical of communities that dwell on soft bottoms where there is high hydrodynamic activity (Pinedo et al., 1997; Sarda` et al., 1999, 2000). No significant variations attributable to the brine discharges from the desalination plant were found. The failure to record any impact may be explained by the high natural variability that is a characteristic feature of bottoms of this type and also by the rapid dilution undergone by the hypersaline brine upon leaving the discharge pipe, inasmuch as the brine does not ordinarily penetrate further than 10 m from the diffuser pipe. This rate of dilution was much faster than the rates that have been reported for discharge pipes having a single outlet instead of diffusers, where the effects of the hypersaline brine discharges have been observed out to a distance of up to 20 m from the outlet (Perez-Talavera and Quesada Ruiz, 2001). Very few studies have assessed the possible effects of brine discharges on the surrounding area, and those that do generally do not compare communities before and after the disturbance. Perez-Talavera and Quesada-Ruiz (2001) examined the effects of a reverseosmosis desalination plant on Cymodocea nodosa and Caulerpa prolifera meadows off

Source of variation

d.f

Pagurus excavatus MS

Fvs

B T(B) L I C B·L B·I B·C T(B) · L T(Bef) · L T(Bef) · I T(Bef) · C T(Aft) · L T(Aft) · I

1 22 2 1 1 2 1 1 44 22 11 11 22 11

3780.88 540.34 5793.24 3764.33 7822.14 48.61 34.76 62.45 402.67 604.45 845.18 363.71 200.90 214.39

No test

T(Aft) · C

11

187.41

Residual

1080

Total

1151

18.42

Anapagurus alboraniensis F

Residual

21.861

T(Aft) · C T(Bef) · I Residual T(Bef) · C

1.144 0.254 10.175 0.515

p

***

***

MS

Fvs

424.13 298.06 5695.18 1841.84 9548.52 2152.51 2244.39 2060.63 259.76 165.27 285.61 44.94 354.25 253.58

No test

454.92 10.56

Anapagurus petiti F

Residual

24.588

T(Aft) · C T(Bef) · I Residual T(Bef) · C

0.557 0.888 43.061 10.123

p

***

***

MS

Fvs

68.06 7.35 18.53 4.97 32.10 3.74 6.78 0.69 8.24 15.95 23.56 8.33 0.53 0.34

No test

0.71

***

Residual

T(Aft) · C T(Bef) · I Residual T(Bef) · C

F

p

23.295

***

0.485 0.015 2.001 0.085

*

(2) (3) (1) (4)

0.35

N. Raventos et al. / Marine Environmental Research 62 (2006) 1–14

Table 2 Asymmetrical ANOVAs comparing mean density of individuals of each selected species at 1 putatively location (D) and 2 controls (C1 and C2)

(continued on next page)

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8

Table 2 (continued) d.f

Mesochaetopterus spp MS

Fvs No test

B T(B) L I C B·L B·I B·C T(B) · L T(Bef) · L T(Bef) · I T(Bef) · C T(Aft) · L T(Aft) · I

1 22 2 1 1 2 1 1 44 22 11 11 22 11

107.56 136.71 693.01 60.39 1325.63 156.05 56.56 255.53 58.75 36.00 54.00 17.99 81.51 57.17

T(Aft) · C

11

105.84

Residual

1080

Total Significance levels:

Trachinus araneus F

Residual

14.112

T(Aft) · C T(Bef) · I Residual T(Bef) · C

0.540 1.059 25.423 5.885

4.16

p

***

*** **

MS

Fvs

1.46 4.06 37.68 22.76 52.61 3.89 0.05 7.72 3.34 3.26 5.21 1.31 3.42 2.03

No test

4.82 T(Bef) · C 0.78

Residual

T(Aft) · C T(Bef) · I Residual 3.671

Xyrichthys novacula F

p

4.300 7.60 2.62 12.59 10.92 0.421 0.389 6.199

< .05;

**p

< 0.01;

***p

***

Fvs

0.38 26.09 54.57 17.02 92.13 26.95 45.56 8.33 8.61

No test

14.53 4.69

*

1151 *p

***

MS

< 0.001. Numbers in brackets refers to statistical test sequence.

1.42

F

p

Residual

6.058

***

T(Aft) · C T(Bef) · I Residual T(Bef) · C

1.781 2.477 4.828 0.658

***

(2) (3) (1) (4)

N. Raventos et al. / Marine Environmental Research 62 (2006) 1–14

Source of variation

N. Raventos et al. / Marine Environmental Research 62 (2006) 1–14 Pagurus excavatus

9

Mesochaetopterus sp.

30

16 14

25

12 20 10 15

8

10

6 4

5 2 0

0

-5

-2 1

3 2

5 4

7 6

9 8

11 10

13 12

15 14

17 16

19 18

21 20

23 22

1 24

3 2

5

7

4

9

6

8

11 10

Anapagurus alboraniensis

13 12

15 14

17 16

19 18

21 20

23 22

24

Trachinus araneus 3.5

45 40

3.0

35 2.5 30 2.0

25

1.5

20 15

1.0

10 0.5 5 0.0

0 -5

-0.5 1

3 2

5 4

7 6

9 8

11 10

13 12

15 14

17 16

19 18

21 20

23 22

1 24

3 2

5 4

7 6

9 8

Anapagurus petiti

11 10

13 12

15 14

17 16

19 21 23 18 20 22 24

Xyrichthys novacula

6

6

5

5

4

4

3

3

2

2

1

1

0

0

-1

-1 1

3 2

5 4

7 6

9 8

11 10

13 12

15 14

17 16

19 18

21 20

23 22

1 24

3 2

5 4

7 6

9 8

11 10

13 12

15 14

17 16

19 18

21 20

23 22

24

Fig. 2. Monthly mean (±0.95 confidence intervals) density of selected species at the putatively impacted location (D = ) and control (C1 = h and C2 = n) locations. Numbers 1–12 and 13–24 indicates before and after periods the plant had begun operating.



10

Source of variation

df

Species richness MS

Fvs No test

B T(B) L I C B·L B·I B·C T(B) · L T(Bef) · L T(Bef) · I T(Bef) · C T(Aft) · L T(Aft) · I

1 22 2 1 1 2 1 1 44 22 11 11 22 11

532.20 32.95 320.59 91.04 550.13 21.22 42.25 0.19 8.43 9.78 7.01 12.55 7.07 6.62

T(Aft) · C

11

7.52

Residual

1080

1.68

Total

1151

Significance levels: *p < .05;

**p

< 0.01;

Shannon–Whiener index F

Residual

5.003

T(Aft) · C

0.881

Residual

4.464

p

***

***

MS

Fvs

3.77 2.48 6.91 1.17 12.65 1.17 1.58 0.76 1.19 1.61 1.29 1.93 0.77 0.51

No test

1.03

Residual

T(Aft) · C T(Bef) · I Residual T(Bef) · C

F

Total abundance p

1.159

5.152 0.395 10.279 0.534

***

***

0.10

***p

< 0.001. Numbers in brackets refers to statistical test sequence.

(2) (3) (1) (4)

MS

Fvs

277047.00 72951.82 136765.00 55942.00 217588.00 29701.00 6152.00 53250.00 40836.73 80821.73 36464.54 125178.91 851.73 614.81 1088.64 755.83

No test

755.83

F

p

B·C Residual Residual

0.116 70.452 37.512

Residual

0.813

(2)

Residual

1.440

(1)

***

(4) (3)

***

N. Raventos et al. / Marine Environmental Research 62 (2006) 1–14

Table 3 Asymmetrical ANOVAs comparing species richness, Shannon–Whiener index and total abundance of individuals at 1 putatively location (D) and 2 controls (C1 and C2)

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Species richness 2.6 2.4 2.2 2.0 1.8 1.6 1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0 0 11 12 13 14 15 16 17 18 19 20 21 22 23 24 1 2 3 4 5 6 7 8 9 10 Number of individuals 600

500

400

300

200

100

0 0 21 22 23 24 1 2 3 4 5 6 7 8 9 10 0 11 12 13 14 15 16 17 18 19 20 Shannon-Whiener Index 2.2 2.0 1.8 1.6 1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0 1 2 3 4 5 6 7

8

9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24

Fig. 3. Monthly mean (±0.95 confidence intervals) of community descriptors at the putatively impacted location (D = ) and control (C1 = h and C2 = n) locations. Numbers 1–12 and 13–24 indicates before and after periods the plant had begun operating.



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the Canary Islands and did not observe any marked harmful impact. On the other hand, a comprehensive study of a site in the Western Mediterranean found lower growth and higher mortality rates for Posidonia oceanica at salinity levels above 39.1 psu (Buceta et al., 2003). A number of workers (e.g., Lewis et al., 2002; Guidetti et al., 2003) have pointed out the need to use univariate and multivariate analyses in order to be able to properly detect environmental impacts. In this study these methods of analysis failed to disclose any significant effect ascribable to any such impact, even though care was taken to ensure adequate temporal replication (Underwood, 1997). It could be that the area sampled might be too small to reflect large spatiotemporal fluctuations affecting the communities located there (Benedetti-Cecchi, 2001). However, the individual abundance values for the different taxa (e.g., echinoderms, molluscs) were within the ranges previously described for this same area by other researchers (Pinedo et al., 1996), and the seasonal variations observed for the macroinfaunal community appear to be consistent with those previously described (Pinedo et al., 1996; Sarda` et al., 2000). High variability appears to be intrinsic to this system, in which most organisms follow a contagious distribution pattern, thereby giving rise to high spatial disparities that occur together with pronounced seasonal and interannual variation. The magnitude of these natural variations is sufficiently large to be able to mask possible alterations caused by the disturbance, and in any case any such alterations stayed within the range of the system’s own natural variability. Hypersaline brine discharges tend to affect juvenile fish more than adult fish (Walsh et al., 1998) and thus to exert an influence on the distribution of nursery areas (VegaCendejas and Hernandez-Santillana, 2004). The settlement patterns of certain fish species, e.g., Xyrichtis novacula, Bothus podas and Trachinus araneus, did not display any significant differences between the zone in which the brine was discharged and the control sites, and recruits of these species were observed in the vicinity of the discharge pipe. Possible effects like those that have been reported for other species, for instance, impaired osmoregulatory ability (Schlenk et al., 2003), did not seem to affect the above-mentioned species, most likely on account of rapid dilution of the hypersaline brine in the water. The frequent sightings of fishes such as Conger conger, Diplodus vulgaris, Mullus surmuletus, Chromis chromis, and Serranus cabrilla (unpublished data) above the outlet portion of the discharge pipe indicate that the pipe acted as an attractor for fish in a manner similar to the action of artificial reefs and other underwater structures (Grossman et al., 1997), with no apparent effects of the hypersaline brine being observable. The low impact of the brine on the benthic community contrasts with the effects of other discharges [e.g., sewage outfall (Bell and Harmelin-Vivien, 1982; Smith et al., 1999; Guidetti et al., 2003)], which have been reported to result in significant changes in the number and abundance of species in the area of the disturbance as compared to other, non-impacted areas. The results of the present study do not necessarily mean that the hypersaline brine discharges have no direct effects on the populations present but only that any such effects cannot be discerned in a statistically significant manner on a short-term basis. The absence of any observed impact could be the result of species mobility, the case of most of the species sighted on the visual censuses, or of the small surface area impacted by the discharges (a maximum of 3500 m2). Still, no apparent effects were observed for certain other, sessile species like the polychaete Mesochetopterus rogeri. Consequently, if the number of desali-

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