Trace metal partitioning in Thalassia testudinum and sediments in the Lower Laguna Madre, Texas

Environment International 31 (2005) 15 – 24 www.elsevier.com/locate/envint

Trace metal partitioning in Thalassia testudinum and sediments in the Lower Laguna Madre, Texas Thomas Whelan III *, Jorge Espinoza, Xiomara Villarreal, Maria CottaGoma Department of Chemistry, University of Texas-Pan American, Edinburg, TX 78541, USA Received 16 January 2004; accepted 28 May 2004 Available online 9 August 2004

Abstract Seagrass communities dominate the Laguna Madre, which accounts for 25% of the coastal region of Texas. Seagrasses are essential to the health of the Laguna Madre (LM) and have experienced an overall decline in coverage in the Lower Laguna Madre (LLM) since 1967. Little is known on the existing environmental status of the LLM. This study focuses on the trace metal chemistry of four micronutrient metals, Fe, Mn, Cu, and Zn, and two non-essential metals, Pb and As, in the globally important seagrass Thalassia testudinum. Seasonal trends show that concentrations of most essential trace metals increase in the tissue during the summer months. With the exception of (1) Cu in the vertical shoot and root, and (2) Mn in the roots, no significant positive correlation exists between the rhizosphere sediment and T. testudinum tissue. Iron indicates a negative correlation between the morphological units and the rhizosphere sediments. No other significant relationship was found between the sediments and the T. testudinum tissue. Mn was enriched up to 10-fold in the leaf tissue relative to the other morphological units and also enriched relative to the rhizosphere sediments. Both Cu and Mn appear to be enriched in leaf tissue compared to other morphological units and also enriched relative to the Cu and Mn in the rhizoshpere sediments. Sediments cores taken in barren areas were slightly elevated in Zn relative to the rhizosphere sediments, whereas no other metals showed statistical differences between barren sediment cores and rhizosphere sediments. However, no correlation was measured in T. testudinum tissue and Zn in rhizosphere sediments. Previous studies suggested that Fe/Mn ratios could indicate differences between seagrass environments. Our results indicate that there is an influence from the Rio Grande in the Fe/Mn signature in sediments, and that ratio is not reflected in the T. testudinum tissue. The results from this study show that the LLM contains trace metal concentrations less than or equal to values for uncontaminated locations worldwide. In addition, there appears to be a complex partitioning in the trace metals in the morphological units of T. testudinum tissue and that analysis only of the leaf may not be indicative of the trace metal levels in this important seagrass species. D 2004 Elsevier Ltd. All rights reserved. Keywords: Trace metal partitioning; Thalassia testudinum; Sediments

1. Introduction It is well known that seagrasses are critical to the structure and function of many marine ecosystems (Klumpp et al., 1989). Seagrasses provide habitat, sediment stability, nutrients, and a food source. In addition, annual production of carbon from seagrasses reaches 4  103 g C/m2, ranking them as the most productive submerged habitats (McRoy and McMilan, 1977). Only salt marshes rank above seagrasses in annual carbon production (Maurinucci, 1982).

* Corresponding author. Tel.: +1-956-381-3371. E-mail address: [email protected] (T. Whelan). 0160-4120/$ - see front matter D 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.envint.2004.05.010

Over 50% of the inshore finfish catch is supported by seagrass habitats (Hedgepeth, 1967). However, many of our seagrass ecosystems are declining. For example, Botany Bay Australia showed a 58% decline in coverage of Posidonia australius (Larkum and West, 1990). In Florida Bay, seagrass standing crop has declined up to 93% for Syringodium filiforme and 28% for Thalassia testudinum (Halk et al., 1999; Zieman et al., 1999). Quammen and Onuf (1993) and Onuf (2003, personal communication) report a 450% increase in the unvegetated bay bottom in LLM since 1965. These events are disturbing and dictate the need to better understand the causes of seagrass decline. In Texas, over 25% of the coastal environment is dominated by seagrass based ecosystems (Onuf, 1995).

16

T. Whelan III et al. / Environment International 31 (2005) 15–24

Because of the enclosed nature of the entire LM, constituents are recycled in the form of detritus (Pulich et al., 1976). This makes the LM especially sensitive and vulnerable to input of pollutants since the tidal flushing and freshwater input, characteristic of traditional estuaries, is highly restricted. The residence time and exposure of pollutants, such as trace metals, to the estuary is consequently quite high. Trace metals, unlike pesticides, petroleum hydrocarbons, and acid rain, occur naturally in organisms. In fact, selected trace metals are essential micronutrients and occur in soils and seawater as part of weathering of crustal rocks and marine sediments. High concentrations of certain essential trace metals can however be toxic (Ralph and Burchett, 1998; Prange and Dennison, 2000). Human activities such as maintenance dredging, transportation activities from the Brownsville Ship Channel and Gulf Intracoastal Waterway (GIWW), and runoff from the Rio Grande and the Arroyo Colorado are all potential sources for heavy metals. Up stream on the Arroyo Colorado new wastewater treatment plants have been built to keep up with the burgeoning population, causing potential increases in toxins and nutrients introduced into the enclosed LLM. The Rio Grande Valley area bordering the LLM has been a major agricultural producer for the last 75 years. Agricultural chemicals, including arsenic, were used until 1970 on cotton to defoliate leaves prior to harvesting the cotton. Lead was widely used in fuels and other agricultural chemicals until 1990. Because of low rainfall and corresponding small volumes of freshwater runoff, chances of chronic contaminant input into the LLM are small. However, isolated values were reported for elevated metals in the Brownsville Ship Channel, the Port Isabel turning basin, and in sediments from the Port Mansfield harbor (Bowles, 1983, 1984; Webster, 1986). In addition, background levels of As, Pb, Zn, and Cu (among other metals), were reported in sediments from the mouth of the Arroyo Colorado (TNRCC, 1995). In the northern section of the Upper Laguna Madre (ULM), sediments showed enriched levels of Ba, V, Cu, Zn, Mn, and Fe (Sharma et al., 1999). Values decreased to background levels a few kilometers south. The seagrass, Posidonia oceanica, has been studied as a biomarker of trace metal contamination in various parts of the world, especially the Mediterranean coast (Maserti et al., 1988; Sanchiz et al., 1990; Costantini et al., 1991; Catiski and Panayotidis, 1993). Along the US Florida coast, T. testudinum leaves were used to evaluate the As content in several estuaries (Fourqurean and Cia, 2001). In the ULM, several studies of the trace metal content in T. testudinum and Halodule wrightti have been reported (Pulich, 1980). H. wrightti is the most abundant seagrass in the LM system, however T. testudinum is increasing its northward coverage especially in the LLM (Kaldy and Dunton, 1999; Onuf, 1996a,b; Quammen and Onuf, 1993). In addition, there have been recent investigations on sunlight and hydrogen sulfide stress and the resulting physiological responses of T. testu-

dinum in the LLM (Major and Dunton, 2002; Kaldy and Dunton, 1999; Cummings and Zimmerman, 2003; Lee and Dunton, 2000). T. testudinum’s subsediment biomass including the root/rhizome complex can account for over 50% of the total plant biomass (Powell et al., 1989; Fourqurean and Zieman, 1991). Hence, the trace metals in the sedimentary environment could be reflected in the extensive subsediment plant tissue. T. testudinum is a climax species (Zieman, 1982) and is an abundant seagrass in many tropical and subtropical environments worldwide. For these reasons, we have selected T .testudinum as an indicator seagrass species for trace metal determination in the LLM. In this paper, we report concentrations of four important micronutrient metals (Fe, Mn, Cu and Zn) and two nonessential trace elements (As and Pb), in T. testudinum tissues, sediment cores, and rhizoshpere sediments from the LLM. We also evaluate the use of T. testudinum as a biomarker organism for trace metals and the relationship between these metals in sediments and T. testudinum tissue.

2. Methods 2.1. Geographic setting The Laguna Madre of Texas (Fig. 1) is a 185-km shallow lagoon (average depth less than 1 m) that originated during the Holocene eustatic sea level rise. The coastal plain bordering the Gulf of Mexico was subsequently inundated concurrent with barrier island development, forming the Laguna Madre of Texas and Mexico (Morton et al., 1998). The Mexican and Texas sections of the LM make up the largest hypersaline lagoon in the world (Tunnell, 2002). The Texas Laguna Madre is divided into two primary water bodies, the ULM and the LLM by a 20-km section of wind tidal flats, commonly known as ‘‘The Land Cut’’. Distinctive Rio Grande deltaic features such as oxbows lakes, distributary bars, and Chenier plains dominate the western shore of the LLM. The GIWW connects these two sections of the Laguna Madre. The LLM, by definition, begins south of the Land Cut extending 93 km to South Bay, which is the southern-most bay in the United States. It should be emphasized that the entire length of the Upper and Lower Laguna Madre is hypersaline averaging over 40 ppt during most of the year. However, during the last 30 years, hypersalinity has decreased from over 50 ppt by connection to the Gulf of Mexico via the jettied passes at the Mansfield Channel, the Brazos Santiago Pass at Port Isabel, and the GIWW (see Fig. 1). Salinity decrease is considered one of the primary reasons that coverage of T. testudinum is increasing in the LLM (Onuf, 1996a; Quammen and Onuf, 1993). The LM of Texas extends 185 km south from Corpus Christi to the Rio Grande, and covers 25% of the Texas coast. Historically the land bordering the LLM has been primarily used for agriculture. The Arroyo Colorado is a

T. Whelan III et al. / Environment International 31 (2005) 15–24

Fig. 1. Location map showing study area in the Lower Laguna Madre.

northern distributary channel of the Rio Grande, which flows directly into the LM and drains most of the agricultural land in the Rio Grande Valley. Currently over 350,000 people occupy the cities and towns bordering the LLM with a projected 30% increase in population by the year 2010 (LRGVDC, 2002). Both continued use of agricultural chemicals and population growth will increase the environmental stress on the LLM. Currently, over 70% of the floor of the LM is covered by seagrasses (Onuf, 1996a). North of the LM are estuaries and bays containing less extensive seagrass beds covering from 0.3% of the sediment surface in the Galveston Bay system to 12.1% coverage in the Corpus Christi Bay system (Onuf, 1995). Because of the subtropical setting of the LM, where evaporation exceeds rainfall, and its restricted access to the open ocean, the LM is hypersaline. The LM extends south into Mexico another 185 km and is also dominated by hypersalinity and seagrass productivity. 2.2. Field methods Our sampling locations are shown on the map in Fig. 1. Monotypic areas of T. testudinum were selected at four

17

locations in the LLM. Location LMT-050 is close to one of South Padre Island’s water treatment plants in 0.5-m water depth. LMT-051 is northwest about 3 km of LMT-050 but in a very different setting. Location LMT-051 is located in about 1.5 m of water in an area well away from anthropogenic influence. Seawater at this site is clear throughout the year with visibility typically 10 m horizontally. Additional experiments are currently underway at this especially beautiful and pristine site. Further north, T. testudinum beds were located at LMT-055 and LMT-056. These locations are between Mansfield Pass and Arroyo Colorado River in about 1-m water depth. Above and subsediment T. testudinum tissues and associated rhizosphere sediments were collected by hand. Immediately after collection, leaves, roots – rhizomes and vertical shoots were washed in the Laguna water to remove sediments, shells and other debris, placed in plastic bags and stored on ice until processing in the laboratory. Rhizosphere sediments were separated from the root – rhizome complex and also stored in separate plastic bags. The vertical shoots contained tissue found just at the sediment – water interface and included new as well as old tissues. Since the vertical shoot is growing at the sediment – water interface, trapped sediment must be thoroughly removed to eliminate any interference from sediment particles. Sediment cores were collected at the locations shown on the map in Fig. 1. Core tubes were made of PVC (7 cm id by 100 cm long) and were pushed by hand from a boat into the sediment as far as possible, usually about 30 – 60 cm penetration. In the head of the coring apparatus is a one-way valve that allows water to escape during penetration into the sediment but closes on retrieval to create a vacuum and keep the core from slipping out. Core barrels were capped on the bottom and taped at the top in order to keep the core sealed until transport to the laboratory. Sediments from the Rio Grande near the mouth were collected by hand coring from the upper 5 to 10 cm of the river bottom. 2.3. Laboratory methods In the laboratory, seagrass tissue was washed again with bay water to remove residual sediment and other debris. Previous studies showed that rinsing seagrasses with distilled water caused premature leaching of metals and other cations (Ledent et al., 1995). Seagrass tissue was separated into leaf blades, roots/rhizomes and vertical shoots, blotted with paper towels and set out at room temperature until partially dry in the laboratory. Epiphytes were easily removed from the partially dry leaves by scraping. We estimate that about 85% of the epiphytic organisms were removed from the leaf surface. After partial drying at room temperature, all of the tissue was oven dried at 95 jC for 24 h followed by grinding to a powder with a ceramic mortar and pestle. The powdered tissue was stored in sealed plastic 50-ml centrifuge tubes until analysis.

18

T. Whelan III et al. / Environment International 31 (2005) 15–24

Sediment cores were extruded onto a PVC tray by pushing the sediment with a PVC piston just slightly smaller than the internal diameter of core barrel. Once on the tray, the core was split lengthwise into two halves and separated. Two sediment samples, about 25 g each, were taken from the center section of the split core at 3 and 25 cm from the top to ensure a representative sample. All seagrass detritus and shell material were removed from both the sediment cores and rhizosphere sediments prior to drying at 95 jC for 24 h or until dry. Dry samples were gently crushed using a mortar and pestle, and any remaining roots and larger shell material were removed by sieving. Sediments were also stored in 50-ml centrifuge tubes until analysis. About 0.5-g samples of seagrass and sediment were weighed to 0.1 mg and digested using concentrated nitric acid and a hydrogen peroxide treatment according to USEPA SW 846-3050 methodology. This procedure leaches and oxidizes the bioavailable metals and is not intended to be a complete digestion that oxidizes and dissolves bound iron and aluminosilicates. A complete digestion requires treatment with HF and HClO4 to oxidize and dissociate all minerals including quartz and aluminosilicates. Table 1 shows recoveries for the metals reported in this study compared to the complete digestion results for the reference materials. These reference materials were selected to cover the range of concentrations possibly encountered in the sediments and seagrasses analyzed in this work. After Table 1 Recovery of reference materials

Zinc (n = 13) S.D. True value % Recovery Copper (n = 11) S.D. True value % Recovery Manganese (n = 7) S.D. True value % Recovery Iron (n = 12) S.D. True value % Recovery Arsenic (n = 17) S.D. True value % Recovery Lead (n = 18) S.D. True value % Recovery

HOS-2-G

HS-2

2006 200 2500 80 (95) 142 6 160 89 15.9 1 14.0 114 227 17 270 84 7.2 1 9.5 76 (106) 0.64 0 0.70 92

109 15 125 87 26 1 26 99 655 36 700 94 21,407 7788 37,000 58 (99) 8.8 1 9.5 92 26 3 32 82

The recovery number in parentheses is recovery from lab using USEPA SW-846 method 3050 and ICP-OES to determine these reference material values. n = number of replicates and true values refer to the complete HF and HClO4 digestion that is intended to release all trace metals from the matrix.

digestion, the samples were filtered and transferred into 50-ml volumetric flasks, brought to volume, and transferred into 50-ml centrifuge tubes for analysis. All the samples reported here were analyzed for Cu, Zn, Fe, and Mn using a PE Analyst 800 flame atomic absorption instrument equipped with a high efficiency nebulizer. Samples were aspirated in triplicate and if the relative percent standard deviation was greater than 5%, samples were re-analyzed. Arsenic and Pb were analyzed by graphite furnace atomic absorption on the same instrument equipped with Zeeman background correction. Each graphite furnace sample is analyzed in duplicate, and if the relative standard deviation is greater than 10%, the sample is re-analyzed. Working standard solutions were made from dilutions of each element from 1000 ppm NIST traceable standard solutions (Fisher Scientific and Spectrum). Quality assurance samples were analyzed in each analytical batch or every 20 samples. Matrix spikes were run for each matrix type and ranged from 80% to 115% recovery. Replicate analyses were made to determine precision and ranged from 85% to 112% RPD. Method blanks were less than 0.75 Ag/g, with the exception of Fe, which had 3.0 Ag/g blank. Reference materials were analyzed with each analytical batch and are shown in Table 1. The reference materials are from the trace metal laboratory at Texas A&M University where (a) HS-2 is a composite marine sediment from the Mississippi River delta that has been freeze dried, mixed and ground to less than 63 Am and (b) HOS-2B is a composite oyster tissue from Galveston Bay that was treated in the same manner. Recovery of these reference materials is one of the most important indicators of data quality.

3. Results In order to better understand the relationships between the three primary morphological units of T. testudinum tissue, samples were collected at all four locations in July 2002 during the growing season and in December 2002 when seagrasses were dormant. The seasonal variations between T. testudinum tissues are shown in Table 2. The percent differences are calculated by subtracting the summer from the winter metal concentrations, dividing by the summer values and multiplying by 100. Negative values for the seasonal data indicate that the summer values are higher than the winter values. Studies in controlled microcosms of T. testudinum showed an increase in accumulation of radioactive labeled 22-Na and 137-Cs with increasing temperatures up to 32 jC (Schroeder and Thorhaug, 1980). These experiments suggest that other non-toxic metals may also show an increase accumulation during the summer months. The 40-year average air temperature difference between July (28.7 jC) and December (16.6 jC) in Brownsville, Texas is 12.1 jC (Tunnell, 2002). Brownsville is about 65 km southwest of LMT-055 (our northern-most location) and should have the same average temperatures. Thus, the

T. Whelan III et al. / Environment International 31 (2005) 15–24 Table 2 Seasonal variations in morphological units of T. testudinum in the Lower Laguna Madre Summer

Winter

12.1 F 3.1 24.1 F 5.7 256 F 96 287 F 77 1.04 F 0.2 0.81 F 0.1

7.1 F 0.5 24.6 F 5.7 95 F 69 169 F 49 1.97 F 0.3 1.29 F 0.4

41 2 63 42 89 59

Vertical shoot Cu 14.0 F 9.7 Zn 34.0 F 12 Mn 71.5 F 51 Fe 541 F 583 As 1.62 F 0.2 Pb 1.17 F 0.3

7.08 F 1.5 40.2 F 13 65.5 F 12 888 F 653 2.20 F 0.6 1.60 F 0.3

49 18 8 64 36 37

Root/rhizome Cu Zn Mn Fe As Pb

4.90 F 0.3 23.4 F 5.3 21.0 F 16 113 F 38 1.21 F 0.9 1.20 F 0.7

38 8 53 73 18 60

Leaf Cu Zn Mn Fe As Pb

7.96 F 3.8 25.4 F 4.3 45.0 F 17 418 F 84 1.47 F 0.7 0.75 F 0.0

Percent change

These results are based on four measurements. The percent change is calculated from the difference between the summer and winter values divided by the summer values times 100.

mean maximum water temperature difference between summer highs and winter low should not exceed 12.1 jC. In addition to lower winter water temperature, light penetration and duration are also reduced during the winter (Major and Dunton, 2002), slowing the growth of T. testudinum. Copper shows a winter decrease in all three morphological units from 38% to 49%, whereas Zn is variable ranging from an 18% increase for the vertical shoots, 2% in the leaves, and 8% in the root/rhizomes. Seasonal variations in Fe and Mn show similar trends for leaf and root/rhizome tissue. Leaf tissue decreases 63% and 42% during the winter for Mn and Fe, respectively. Root/ rhizome tissue decreased 53% and 73% for Mn and Fe, respectively. The vertical shoots for Mn and Fe are more complex. Fe substantially increases in the winter, whereas the Mn decreases slightly by 8%. Mn and Fe are both important in plant metabolism and are required in the photosynthetic processes occurring in the leaves that increase during the high growing period in summer months. Because As and Pb concentrations are small and are near the limit of detection, seasonal changes for these metals are quite large. However, with one exception for As (root/ rhizomes), winter values both Pb and As are higher than for summer. The important result here is the trend of lower T. testudinum tissue values in summer for both As and Pb. As suggested by others (Schlacher-Hoenlinger and Schlacher, 1998), this is probably a result of summer growth ‘‘diluting’’ Pb and As which are non-essential elements for seagrass growth and are incorporated into leaf tissue by

19

passive uptake rather than active metabolic incorporation inside the cellular structure. The histograms in Figs. 2 –4 show the results for trace metals in the morphological units of T. testudinum at four locations in the LLM. Fig. 3 shows Fe plotted separately for the sediment and tissue to better illustrate the differences between tissue and rhizoshpere sediments. Some general observations are evident. First, as indicated by the results shown in Figs. 2 – 4, rhizosphere sediment at LMT-055 contained the highest values for all six metals of the four sites where the seagrass was studied. However, this location did not contain the highest metal values for the T. testudinum tissues. Additionally, at LMT-050, the seagrass leaf tissue contained more Cu, Zn, and Mn than these metals found in the associated rhizoshpere sediments. This suggests that even if all the acid leachable metal in the sediment was available to the plant, it could not account for the metal concentration contained in the leaf tissue. Examination of Mn shows that the leaf tissue in three of the four locations contain more Mn than the available rhizosphere sediments. This reinforces the probability that metals in the seagrass leaf tissue, unlike land plants, do not come from the sediments. The data in Fig. 3 shows Fe plotted with just the T. testudinum tissue (bottom plot) to illustrate the differences just the morphological units of T. testudinum. As opposed to the Mn results, the Fe data is more complex. First, the rhizosphere sediments contain about 10 to 20 times more Fe than any of the morphological tissue units of T. testudinum. The root/rhizomes have relatively consistent values near 400 Ag/g for all sites. LMT-055 showed the highest Fe in the

Fig. 2. Distribution of Copper and Zinc for rhizosphere sediment, leaf, vertical shoot, and root/rhizome system at four locations in the LLM.

20

T. Whelan III et al. / Environment International 31 (2005) 15–24

Fig. 4. Distribution of Aresnic and Lead for rhizosphere sediment, leaf, vertical shoot, and root/rhizome system at four locations in the LLM.

The range for Pb in seagrass tissue is < 0.75– 1.53 Ag/g and for As, < 0.75 – 2.2 Ag/g. Arsenic is within the lower range reported for T. testudinum leaves in Florida estuaries (Fourqurean and Cia, 2001). Very little information is

Fig. 3. Distribution of Iron and Manganese for rhizosphere sediment, leaf, vertical shoot, and root/rhizome system at four locations in the LLM.

rhizosphere sediment but contained close to the lowest value for the leaf. However, the leaf tissue in LMT-051 contained the most Fe yet had the least amount of Fe in the rhizosphere sediments. In Table 3, results for correlation coefficients between the T. testudinum tissue and the rhizosphere sediments are also revealing. With the exception of Cu in the vertical shoot, the root/rhizome, and the Mn root, r2 values are generally less than 0.5. The correlation coefficients for Fe are all negative. These data further indicate that the source of metals in the T. testudinum leaf tissue is not derived from the rhizoshpere sediments. In contrast to the other metals, Mn showed the highest values in the leaf. Table 3 shows mean values for roots/rhizomes and vertical shoots were five times less than Mn in the leaf. With the exception of Mn in LMT-055, the rhizosphere sediments contained less Mn than the leaf. This result suggests that Mn bioaccumulates in leaf tissues of T. testudinum in the LLM. The vertical shoot results show the most consistent values for Mn, with an average value of 44.3 F 4 Ag/g.

Table 3 Average, standard deviation and correlation coefficient between metals in Thalassia tissue and rhizoshpere sediment Metal

Rhizoshpere sediment

Leaf

Vertical shoot

Root

Copper Average F S.D. r2

11.5 F 3

10.3 F 5 0.39

12.9 F 11 0.77

6.4 F 4 0.74

Zinc Average F S.D. r2

32.7 F 11

20.2 F 9 0.13

33.9 F 12 0.007

23.2 F 6 0.34

Manganese Average F S.D. r2

243 F 160

260 F 79 0.49

44.3 F 4 0.37

41.9 F 15 0.80

Iron Average F S.D. r2

9148 F 2639

457 F 357 0.62

710 F 524 0.72

406 F 73 0.40

Arsenic Average F S.D. r2

5.5 F 2.2

1.27 F 0.5 0.65

1.46 F 0.3 0.42

1.39 F 0.7 0.43

Lead Average F S.D. r2

5.2 F 0.8

0.87 F 0.5 0.98

1.02 F 0.5 0.68

0.31 F 0.3 0.10

All units are in Ag/g dry weight.

T. Whelan III et al. / Environment International 31 (2005) 15–24

21

Table 4 Sediment cores for north – south transect in Lower Laguna Madre Field ID

Cu

Zn

Mn

Fe

As

Rio Grande SB6 001 2 cm 25 cm LMT-050 2 cm 25 cm LMT-051 2 cm 25 cm LMT-052 2 cm 25 cm LMT-053 2 cm 25 cm LMT-054 2 cm 25 cm LMT-055 2 cm 25 cm LMT-056 2 cm 25 cm Mean 2 cm Standard deviation Mean 25 cm Standard deviation Upper Lagunaa

13.8

57.5

344

18,000

0.3

13.0 13.1

43.0 52.5

235 235

13,264 13,310

5.3 5.4

10.1 10.1

10.5 12.3

49.5 54.3

95 181

7925 12,385

5.3 6.2

5.4 8.3

9.00 10.9

38.2 54.4

123 207

8350 7875

3.6 5.1

5.6 4.7

13.9 12.0

47.7 55.4

280 281

11,145 17,210

3.0 3.6

5.6 6.5

12.1 10.4

80.9 27.1

405 120

17,350 6390

2.1 9.7

5.3 2.9

Fig. 5. Fe/Mn values for seagrass and rhizosphere sediment, leaf, vertical shoot, and root/rhizome system at four locations in the LLM.

6.90 7.10

39.0 52.8

170 147

8905 8980

3.2 2.8

4.0 4.0

9.10 14.0

48.4 32.1

163 700

6805 14,375

1.6 9.8

4.0 5.0

6.30 3.80 10.1 3

34.5 45.2 47.7 14

134 75 201 102

6790 4150 10,067 3678

3.2 2.8 3.4 1.3

2.8 1.8 5.4 2.2

10.5 3

46.7 11

243 196

10,584 4438

5.7 3

5.4 3

core samples were taken in barren areas where seagrasses were not presently growing and are listed from south to north along the same transect as the seagrass sampling transect. In addition, sediments from South Bay and the lower section of the Rio Grande are reported to compare with the LLM sediment composition. Pulich et al. (1976) suggested Fe/Mn values were useful in distinguishing between two seagrass populations in the ULM system. The data in Table 5 and in Fig. 5 show the Fe/Mn values for all the tissue and the rhizosphere sediment for the LLM in this study. The Fe/Mn value for the sediment is by far the highest. The vertical shoot and root/rhizomes contain the next highest Fe/Mn ratios. The leaf shows the smallest Fe/ Mn ratios. This is indicative of naturally enriched Mn in T. testudinum leaf tissue. Rio Grande sediment was taken from the upper 10 cm of surface sediments. The data in Table 5 indicate the trace metals from cores and rhizosphere sediments taken in the Rio Grande South Bay and LMT-050 (our southern-most location) of the study area show similar Fe/Mn values. This suggests that recent sediments from the Rio Grande dominate the trace metal composition in the southern portion of the LLM. Sediments north of LMT-051 show a decreasing influence from the Rio Grande deltaic sediments. Pulich et al. (1976), however, show considerably smaller Fe/Mn for the ULM. This is most likely a result of the weak acid leaching procedure used in their sample preparation for metal analysis. The more rigor-

< 0.6 – 19 4 – 42 10 – 150 700 – 6700 NA

Pb 8.58

< 8 – 81

a

Sharma et al. (1999). These authors used complete digestion procedures using HF and HClO4.

available for Pb in T. testudinum tissue. Both Pb and As are lower in the rhizosphere sediments than the four micronutrient metals discussed above. In addition, the correlation coefficients for tissue and rhizosphere sediments in As and Pb are less than 0.7 for all samples. Pb has a negative 0.98 r2 value for the leaf to rhizosphere sediment relationship, which provides further evidence that the source of T. testudinum trace metals is not directly from the rhizosphere sediments. Sediment core data given in Table 4 shows trace metal composition for the six metals reported in this study. The

Table 5 Fe/Mn data for sediments and T. testudinum

Sediment Leaf Vertical Shoot Root a

Rio Grande

South Bay

LMT-050

LMT-051

LMT-055

LMT-056

Pulich et al.a

Cores, this studyb

Sharma et al.c

52 na na na

57 na na na

56 2 33 10

45 4 20 12

26 1 9 5

47 1 4 16

5 to 11 1 to 3 na 2 to 28

46.5 na na na

67.4 na na na

Note Pulich et al. (1976) used a less rigorous procedure for leachable metals (hydroxlamine – acetic acid). This result is from both sections of eight cores reported in this study. c Sharma et al. (1999) used complete digestion with HF and HClO4 for their sample determinations. b

22

T. Whelan III et al. / Environment International 31 (2005) 15–24

ous procedure used in this study would have leached more Fe than the hydroxylamine-acetic acid procedure. Pulich et al. (1976) reported for their metal results. In contrast, the data in the ULM (Sharma et al., 1999) show a larger value for Fe/Mn than reported here. This is likely a result of more effective Fe release resulting from the complete (HF-HClO4) sample digestion used for the sediments reported in the ULM. It is essential for comparison of sediment data, that the sample preparation and digestion procedures be clearly defined.

4. Discussion The limited source of freshwater and restricted flow of seawater into and out of the LLM (Fig. 1) indicates that only a fraction of the metals and other nutrients come from renewed seawater or freshwater sources. Thus, the recycling of nutrients (and potential pollutants) from seagrass into the LLM must account for much of the metabolic requirements of ecosystem (Capone and Tayor, 1980). Often large rafts of seagrass leaves are found stranded along the bars and islands in the LLM, indicating that the productivity of the seagrass community can move out of the vast live meadows onto other unvegetated areas. This is opposed to the typical estuary where fluvial discharge provides a continual input of nutrients and also gives the estuary a discharge route, exiting into the ocean. This condition makes the entire LM system especially susceptible and sensitive to input of pollutants since residence times for seawater movement into and out of the bay are considerably longer. Natural variations in trace metals occur in T. testudinum within the hypersaline LM system. These variations occur between sediments and morphological units of the seagrass, as well as season. The concentration variations and ranges of trace metals in Laguna sediments and T. testudinum reported here are well within values for uncontaminated sites (Campanella et al., 2001; Pulich et al., 1976; Sharma et al., 1999). However, in the event of a discharge of pollutants, such as trace metals, into the LM system, seagrass and organisms depending upon seagrass, will be in contact with the pollutant longer than for traditional estuaries. Seagrasses, unlike land plants, may not necessarily acquire micronutrient metals from sediments. Seawater could provide the source of trace metals required for growth and development of seagrasses. The trend in this study and others show that most of the metals in the leaf and vertical shoot tissue do not come from the sediment. This is in opposition to freshwater aquatic plants (Jackson et al., 1991; Jackson, 1998) where metals can be transported from freshwater sediments into the leaves, roots and vertical shoots. Schroeder and Thorhaug (1980) showed that radioactive-labeled metals Cd, Mn, Zn and Cs added to seawater in controlled microcosms were taken up by the leaves of T. testudinum and were significantly excluded from the roots and rhizomes. Leaf material in Zostera marina showed that following uptake of radioactively labeled Cd and Mn treated

seawater, a rapid release of Cd was observed but Mn was slowly released back into the seawater (Brinkhuis et al., 1980). This is presumably because Mn is an essential nutrient and Cd is not. Micronutrient trace metal content in seagrasses is generally higher in summer than winter. This is consistent with the trend in seagrasses found in the ULM (Pulich et al., 1976), and in the Mediterranean (Campanella et al., 2001) for P. oceanica. T. testudinum is complex in micronutrient metal distribution and appears to vary in concentration for each metal and morphological unit. These results suggest that if T. testudinum is to be used as a biomonitor, care must be taken to analyze all the morphological units and, depending upon the metal(s) of concern, the season of maximum concentration should be considered. For example, Mn, Fe and Cu have maximum concentrations in the summer, whereas Pb and As have maximum concentrations in the winter. Previous investigations reported trace metals in sediments (Schiff and Weisberg, 1999) and seagrasses (Pulich et al., 1976) normalized to the Fe content indicate both source and anomalous concentrations of other metals. According to the Fe data for the LLM reported here, correlation between the tissue and sediment is either random or inversely related (Table 3). The variability in micronutrient metal content between the morphological units is highly dependent upon the metal in question. For example, Mn is highest in the leaf tissue by a factor of 3 to 10 over the vertical shoot and the root/ rhizomes. In fact, in three of the four locations, leaf tissue contains more Mn than the associated rhizoshpere sediment. This is in contrast to Zn where concentrations decrease from the vertical shoot, to the root – rhizome, to the leaf. Fe shows a large variation within the leaf and vertical shoot, whereas the root is more consistent between samples. Fortunately, Fe is normally not toxic to seagrasses, even at elevated concentrations, hence the variation in concentration between morphological units is not of considerable importance for Fe. The usefulness of Fe/Mn values in seagrasses as an environmental indicator is not apparent in the LLM.

5. Conclusions 1. The results from this study indicate that trace metals examined in T. testudinum are fractionated within morphological units and between the rhizosphere sediments. In most cases, Cu and Mn demonstrated higher values in leaf tissue than the rhizosphere sediments. The implication is that these metals enter the plant leaf from the water column rather than from the sediments. 2. Rhizosphere sediments, cores taken in barren areas and T. testudinum tissue metal concentrations in the LLM are within the range of uncontaminated estuaries. 3. Sedimentary Fe/Mn values used in other estuaries as environmental indicators do not appear relevant in T. testudinum signatures in the LLM.

T. Whelan III et al. / Environment International 31 (2005) 15–24

4. If T. testudinum is used a biomonitor for trace metals in contaminated sites, both season and morphological unit must be selected prior to sampling, depending upon the metal in question.

Acknowledgements We wish to thank the University of Texas-Pan American Faculty Research Council and the Robert A. Welch Foundation for their support of this work. Special thanks to Mr. Donald Hockaday, director of the UTPA Coastal Studies Laboratory, for his field support for this project, and Dr. Hudson DeYoe for the helpful discussions on this manuscript. This is part of the Center for Subtropical Studies series CSS 2004-001. We also wish to thank the anonymous reviewers for their recommendations for improvement of this manuscript.

References Bowles Jr WF. Report to the Texas Department of Water Resources. Intensive survey of the Brownsville ship channel. Field Measurements, Water Chemistry, Sediment Chemistry and Biological Indices, vol. 2494. Austin (TX): Segment; 1983. p. 1 – 49. Bowles Jr WF. Report to the Texas Department of Water Resources. Intensive survey of Port Mansfield measurements, water chemistry, sediment chemistry and biological indices. Texas Department of Water Resources. 1984;15 – 68 [Austin (TX)]. Brinkhuis BH, Penello WF, Churchill AC. Cadmium and manganese flux in eelgrass Zostera marina: II. Metal uptake by leaf and root – rhizome tissues. Marine Biology 1980;58:187 – 96. Campanella L, Conti ME, Cubadda F, Sucapane C. Trace metals in seagrasses, algae and mollusks from an uncontaminated area in the Mediterranean. Environmental Pollution 2001;111:117 – 26. Capone DG, Tayor BF. Microbial nitrogen cycling in a seagrass community. Estuarine Perspectives. 1980;153 – 61. Catiski VA, Panayotidis P. Copper, chromium and nickel in tissues of the Mediterranean seagrasses Posidonia oceanica and Cymodocea nodosa (Potamoetonaceae) from Greek coastal areas. Chemosphere 1993;26: 963 – 78. Costantini S, Giordano R, Ciaralli L, Becaloni E. Mercury, cadmium and lead evaluation in Posidonia oceanica and Codium tomentosum. Marine Pollution Bulletin 1991;22:362 – 3. Cummings ME, Zimmerman RC. Light harvesting and the package effect in the seagrasses Thalassia testudinum Banks ex Ko¨nig and Zostera marina L.: optical constraints on photoacclimation. Aquatic Botany 2003;75:261 – 74. Fourqurean JW, Cia Y. Arsenic and phosphorus in seagrass leaves from the Gulf of Mexico. Aquatic Botany 2001;71:247 – 58. Fourqurean JW, Zieman JC. Photosynthesis, respiration and whole plant carbon budget of the seagrass Thalassia testudinum. Marine Ecology Progress Series 1991;69:161 – 70. Halk MO, Durako MJ, Fourqurean JW, Zieman JW. Decadal changes in seagrass distribution and abundance in Florida Bay. Estuaries 1999;22: 445 – 59. Hedgepeth JW. Ecological aspects of the Laguna Madre, a hypersaline estuary. ‘‘Estuaries’’. Washington, DC: American Association for the Advancement of Science, 1967. pp. 408 – 19. Jackson LJ. Paradigms of metal accumulation in rooted aquatic vascular plants. Science of the Total Environment 1998;219:223 – 31.

23

Jackson LJ, Rasmussen JB, Peters RH, Kalff J. Empirical relationships between the element composition of aquatic macrophytes and their underlying sediments. Biogeochemistry 1991;12:71 – 96. Kaldy JE, Dunton KH. Ontogenetic photosynthetic changes, dispersal and survival of Thalassia testudinum (turtle grass) seedlings in a sub-tropical lagoon. Journal of Experimental Marine Biology and Ecology 1999;240:193 – 212. Klumpp DW, Howard RK, Pollard DA. Trophodynamics and nutritional ecology of seagrass communities. Biology of seagrasses. Amsterdam: Elsevier; 1989. p. 394 – 457. Larkum AWD, West RJ. Long-term changes of seagrasses meadows in Botany Bay, Australia. Aquatic Botany 1990;37:55 – 70. Ledent G, Mateo MA, Warnau M, Temara A, Romero J, Dubois PH. Element losses following distilled water rinsing of leaves of the seagrass Posidonia oceanica (L) Delile. Aquatic Botany 1995;52: 229 – 35. Lee KS, Dunton KH. Diurnal changes in pore water sulfide concentrations in the seagrass Thalassia testudinum beds: the effects of seagrasses on sulfide dynamics. Journal of Experimental Marine Biology and Ecology 2000;255:201 – 14. LRGVDC: Lower Rio Grande Valley Development Council. 2002 Annual report. Major KM, Dunton KH. Variations in light harvesting characteristics of the seagrass, Thalassia testudinum: evidence for photoacclimation. Journal of Experimental Marine Biology and Ecology 2002;275:173 – 89. Maserti BE, Ferrar R, Paterno P. Posidonia as an indictor of mercury contamination. Marine Pollution Bulletin 1988;19:381 – 2. Maurinucci AC. Trophic importance of Spartina alterniflora production and decomposition to the marsh estuarine ecosystem. Biological Conservation 1982;22:35 – 58. McRoy CP, McMilan C. Production ecology and physiology of seagrasses. Seagrass Ecosystems: A Scientific Perspective. 1977;53 – 87. Morton RA, White WA, Nava RC. Report for the Bureau of Economic Geology. Sediment budget analysis for Laguna Madre, Texas: an examination of sediment characteristics, history, and recent transport. Austin (TX): University of Texas; 1998. p. 194. Onuf CP. Seagrass meadows of the Laguna Madre of Texas. Our living resources: a report to the nation on the distribution, abundance, and health of US plants, animals and ecosystems. Washington (DC): U.S Department of the Interior, National Biological Service; 1995. p. 275 – 7. Onuf CP. Biomas patterns in seagrass meadows of the Laguna Madre, Texas. Bulletin of Marine Science 1996a;58:404 – 20. Onuf CP. Seagrass responses to long-term light reduction by brown tide in upper Laguna Madre, Texas: distribution and biomass patterns. Marine Biology Progress Series 1996b;138:219 – 31. Powell GMV, Kenworthy WJ, Fourqurean JW. Experimental evidence for nutrient limitation of seagrass growth in a tropical estuary with restricted circulation. Bulletin of Marine Science 1989;44:324 – 40. Prange JA, Dennison WC. Physiological response of 5 seagrass species to trace metals. Marine Pollution Bulletin 2000;41:327 – 36. Pulich WM. Heavy metal accumulation by selected Holodule wrightii Asch. populations in the Corpus Christi bay area. Contributions in Marine Science 1980;23:90 – 100. Pulich WM, Barnes S, Parker PL. Trace metal cycles in seagrass communities. Estuarine Processes, ‘‘Uses Stresses, and Adaptation to the Estuary’’ 1976;1:493 – 506. Quammen ML, Onuf CP. Laguna Madre: seagrass changes continue decades after salinity reduction. Estuaries 1993;16:302 – 10. Ralph PJ, Burchett MD. Photosynthetic response of Halophila ovalis to heavy metal stress. Environmental Pollution 1998;103:91 – 101. Sanchiz C, Benedito V, Pastor V, Garcia-Carrascosa AM. Bioaccumulation of heavy metals in Posidonia oceanica (L.) Delile and Cymodocea nodosa (Ucria) Aschers. at an uncontaminated site in the east coast of Spain. Mer Medit 1990;32(3). Sharma VK, Rhudy KB, Koening R, Vazquez FG. Metals in sediments of the Upper Laguna Madre. Marine Bulletin 1999;38:1221 – 6.

24

T. Whelan III et al. / Environment International 31 (2005) 15–24

Schiff KC, Weisberg SB. Iron as a reference element for determining trace metal enrichment in Southern California coastal shelf sediments. Marine Environmental Research 1999;48:161 – 76. Schlacher-Hoenlinger MA, Schlacher TA. Differential accumulation patterns of heavy metals among the dominant macrophytes of Mediterranean seagrass meadows. Chemosphere 1998;37:1511 – 9. Schroeder PB, Thorhaug A. Trace metal cycling in tropical – subtropical estuaries dominated by the seagrass Thalassia testudinum. American Journal of Botany 1980;67:1075 – 88. TNRCC. (Texas Natural Resource Conservation Commission Toxic). Contaminants survey of the Lower Rio Grande, Lower Arroyo Colorado, and associated coastal waters. Austin, TX. AS-69; 1995.

Tunnell JW. Climate and hyrdography. The Laguna Madre of Texas and Tamaulipas. 2002;7 – 27. Webster CF. Report to Texas Water Commission: Port Isabel fishing harbor water quality survey. Physiochemistry, fecal coliform bacteria and benthos. Texas Water Commission. 1986;41 [Austin (TX), (LP 86-09)]. Zieman JC. The ecology of the seagrasses of South Florida: a community profile. Washington, DC: U.S. Fish and Wildlife Service, Office of Biological Services; 1982. p. 158 [FWS/OBS 82/25]. Zieman JC, Fourqurean JW, Frankovich TA. Seagrass die off in Florida Bay (USA): long term trends in abundance and growth of turtle grass, Thalassia testudinum. Estuaries 1999;22:460 – 70.