Microb Ecol (1989) 18:135-146
MICROBIAL ECOLOGY | Springer-Verlag New York Inc. 1989
Phenanthrene Mineralization Along a Natural Salinity Gradient in an Urban Estuary, Boston Harbor, Massachusetts Michael P. Shiaris University of Massachusetts at Boston, Biology Department, Boston, Massachusetts 02125, USA
Abstract. The effect of varying salinity on phenanthrene and glutamate mineralization was examined in sediments along a natural salinity gradient in an urban tidal river. Mineralization was measured by trapping ~4CO2 from sediment slurries dosed with trace levels of [14C]phenanthrene or [14C]glutamate. Sediments from three sites representing three salinity regimes (0, 15, and 30%0) were mixed with filtered column water from each site. Ambient phenanthrene concentrations were also determined to calculate phenanthrene mineralization rates. Rates of phenanthrene mineralization related significantly to increasing salinity along the transect as determined by linear regression analysis. Rates ranged from 1 ng/hour/g dry sediment at the freshwater site to > 16 ng/hour/g dry sediment at the 30%00salinity site. Glutamate mineralization also increased from the freshwater tO the marine site; however, the relationship to salinity was not statistically significant. To examine the effect of salinity on mineralizing activities, individual sediments were mixed with filtered water of the other two sites. Slurries were also made with artificial seawater composed of 0, 15, or 30 g NaCI/ liter to substitute for overlying water. Rates ofphenanthrene mineralization in the 0%0 ambient salinity sediments were not affected by higher salinity waters. Activities in the 15 and 30%0 ambient salinity sediments, however, were significantly inhibited by incubation with 0%0 salinity water. The inhibition, in large part, appears to be due to the decreased NaC1 concentration of the water phase. Glutamate mineralization was affected in a similar manner, but not as dramatically as phenanthrene mineralization. The results suggest that phenanthrene degraders in low salinity estuarine sediments subject to salt water intrusion are tolerant to a wide range of salinities but phenanthrene degradation in brackish waters is mainly a function of obligate marine microorganisms.
Introduction
Polycyclic aromatic hydrocarbons (PAHs), a class of compounds composed of two or more fused benzene rings, are of concern to marine biota and h u m a n public health as toxic, mutagenic, and carcinogenic agents [2, 46]. Phenan-
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threne, a three-ring PAH, is not a mutagen but exhibits toxicity to marine organisms [7, 33]. Compared with higher molecular weight PAHs, phenanthrene is readily biodegraded and serves as a useful model for studying the fate of PAHs in the environment. Individual P A H compounds have low water solubilities [28] which decrease with increasing molecular weight. Consequently, they display a high affinity for organic matter and particles in aquatic ecosystems [31]. With particle settlement, PAHs, including phenanthrene, often reach high levels in aquatic sediments neighboring sources of PAH-generating activities. Although sediments are exposed to naturally derived PAH, it is anthropogenic sources that contribute the predominant share of PAH to urban waters. For example, the depositional history of industrial and energy-generated PAH inputs can be readily chronicled in undisturbed sediments adjacent to urban areas [21]. Urban estuaries, such as Boston Harbor, attain P A H concentrations exceeding 100 mg/g sediment [20, 38, 47]. Once trapped in aquatic sediments, PAHs persist for long periods, and, depending on environmental conditions and the structure of the compound, they may last indeterminately. Chemical oxidation and photooxidation may be important routes of P A H loss in the water column, but biological transformation is probably the prevailing mechanism of P A H loss in sediments. The biological participants include many eukaryotic organisms native to marine sediments [26, 29, 42] as well as a diverse group of bacteria isolated from marine environments [12, 15, 23, 37, 39, 44]. Many of the bacteria, in contrast to the eukaryotes, are capable of utilizing PAHs as sole carbon and energy sources [9]. The importance of microorganisms in the degradation and fate of PAHs in marine sediments has been well documented [4, 17, 18, 25]. The fate of P A H in sediments is under the control of complex biological and environmental factors [1]. For example, oxygen is a master environmental variable that controls degradation rates [11, 16]. Both eukaryotic and prokaryotic degradation pathways utilize bimolecular oxygen in the initial enzymatic attack on the P A H ring [9], yet oxygen is often limiting in sediments. The influx of oxygen into sediments can be augmented by burrowing animals which in turn stimulate microbial degradation of P A H [6, 13]. Furthermore, the degree of mixing of oxygen into anaerobic sediments is influenced by the numbers, types, and seasonal succession of benthic animals in the sediments [27]. In addition to oxygen, P A H biodegradation is affected by nutrients, temperature [1], microbial adaptation to PAHs [35], and the composition of the PAH mixture [5]. In marine environments, salinity also appears to have a significant effect on microbial degradation of PAHs. Kerr and Capone [22] reported that naphthalene and anthracene degraders in the Hudson River estuary were highly adapted to their salinity regime. I observed that degradation of naphthalene and phenanthrene in three urban estuarine sediments was significantly related to seasonal salinity changes over the course of a 15-month period [36]. The same trend was observed for benzo(a)pyrene degradation for one of the sites. These results suggested that P A H degradation was carried out predominantly by obligate marine microorganism. The objective of the present study was to determine if salinity affects phenanthrene mineralization along a natural salinity gradient and whether mineralization is dependent on the presence of sodium chloride.
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GREATER BOSTON AREA
BOSTON HARBOR UMoSs
0------9
50%= SITE
m. ~
0
0,5
!
I
1 IKm
15%. SITE 0%= ;ITE
Fig. I. Map of the Neponset River, a tributary of Boston Harbor, Massachusetts. The three sampling sites are located within the tidal extension of the river. Methods
Sampling Three sites along a transect in the tidal Neponset River were chosen to represent a salinity gradient ranging from 0 to 30%0 (Fig. 1). The Neponset River is a tributary of Boston Harbor, Massachusetts. Boston Harbor is a glacially carved, tidally dominated estuary, typical of many bays in New England. The Harbor and its 31 associated islands cover an area of 114 square kilometers with 190 km of shoreline. All reaches of the Harbor and its tributary streams are polluted (Federal Water Pollution Control Federation, Northeast Water Quality Management Center, 1967, Department of the Interior, Washington DC), and at least one-third of the harbor is grossly polluted by municipal and industrial sewage with deposits of decayed organic matter and oil residues covering much of the bottom (M. G. Fitzgerald, 1980, Woods Hole Oceanographic Institute, WHOI-80-38, Woods Hole, MA). Freshwater influx to the Harbor is low and the entire Harbor is flushed in 42 days, predominantly by diurnal tides. Aside from two main shipping channels, the mean depth of the harbor is 3 m or less at mean low water. All three sites were composed of muds consisting of fine silt-clays. Sediment samples were taken in triplicate at each site with an Eckman dredge (Wildco, Saginaw, Michigan). The top 2 - 3 - m m sediment of each grab was subsampled with a sterile wooden spatula,
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placed in glass bottles, kept at ambient water temperature in chest coolers, and returned to the laboratory for immediate processing. Salinity and conductivity (Beckman RS5-3 Salinometer), sediment temperature, and pH were recorded at each site and sampling date. Total organic carbon by sediment acidification was determined on a C H N analyzer by the Darling Marine Laboratory, Walpole, Maine. Ambient phenanthrene concentrations were determined by Soxhlet extraction of sediments with methanol and methanol/benzene, column chromatography fractionation, and high performance liquid chromatography/diode array detection analysis, as described previously [38].
Phenanthrene Mineralization Rates Techniques for determining PAH mineralization rates were based on the protocol of Herbes and Schwall [19]. Surficial sediment samples, one from each sediment grab, were diluted 1:10 (wt/wt) in estuarine water or autoclaved m i n i m u m salts solution (MSS) and homogenized with a magnetic stirrer. The salinity of the estuarine water or MSS was either 0, 15, or 30%0. MSS was composed of 0.8 g K2HPO4, 0.2 g KH2PO4, 0.05 g CaSO4-2H20, 0.5 g MgSO4-7H20, 1.0 g (NH4)2SO4,0.01 g FeSO4"2H20, and 1 liter distilled water (pH 8.0) and amended with either 0, 15, or 30 g NaC1/ liter final concentration. The estuarine water was obtained from the water column overlying th~ sampling site and filtered through a 0.2-#m Nuclepore membrane prior to mixing. Aliquots of the sediment slurry (10 ml) were transferred via sterile, wide-mouth, serological pipets to sterile 25ml Erlenmeyer flasks, one for each sediment grab, therefore three per site. The flasks were fitted with Teflon-lined serum caps and glass center wells. Aliquots (0.1 ml) ofradiolabeled PAH dissolved in acetone were added to each incubation flask at a final concentration of 20 ng/g wet slurry, a concentration below the ambient PAH concentrations. Flasks were incubated at in situ temperatures in the dark without shaking for 60 hours. During the incubation, sediments settled to the bottom of the flask at an approximate thickness of 1 ram. No discernable change in the light color of the sediments occurred during incubation indicating that the sediments remained oxidized. Mineralization rates were linear with time (data not shown), which also indicated that oxygen was not depleted. The incubation time resulted in less than 10% transformation of the parent radiolabeled phenanthrene. All experiments included frozen controls, which were identical to occasional controls killed by autoclaving or formalin (2% final concentration) addition. Incubations were terminated by injecting 0.3 ml 10% trichloroacetic acid through the Teflon-lined serum cap. Radiolabeled CO2 was trapped in the glass center wells by the addition of 0.3 ml 2 N K O H to the wells. K O H was used rather than an organic solvent CO2-trapping agent to minimize contamination with radiolabeled phenanthrene. Flasks were shaken for 30 m i n at 1 I0 rpm and the K O H was transferred into vials containing 10 ml Aquasol (New England Nuclear, Boston, MA) for liquid scintillation counting. Counts were performed on a Packard Model 3330 Tricarb liquid scintillation counter. Quenching was determined by the internal standards method employing [t4C]glutamate for the Aquasol cocktails. Preliminary experiments were conducted to ensure that the center well KOH was free of [~4C]phenanthrene. The KOH was extracted with benzene and examined by thin-layer chromatography (TLC) and liquid scintillation counting. Phenanthrene could not be detected on TLC and scintillation counts were less than 0.1% of the [14C]phenanthrene counts added to the flasks. Mineralization rates were calculated with the assumption that added radioactive PAH was completely mixed and equilibrated with ambient sediment PAH. The mineralization rate, R, was calculated as follows: R = r (Ca + C~)/Cr where, r = rate of [14C]phenanthrene mineralization corrected for sterile controls Ca = ambient concentration of PAH in sediment C, = concentration of added ~4C-phenanthrene
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Because [~4C]phenanthrene was added in trace amounts compared to known ambient phenanthrene concentrations, R represents an accurate estimate of mineralization rates in the sediment slurries.
Glutamate Mineralization Incubationflasks were prepared as for phenanthrene mineralizationassays except that plastic center wells and rubber septa without Teflon linings were used. Radiolabeled glutamate was added at a final concentration of 130 pg/ml of slurry. Incubations were carded out in the dark at in situ temperatures for 1 hour. Replicate controls consisted of sediment slurries killed with 0.5 ml m-cresol. Incubations were terminated by the addition of 0.3 ml 2 M sulfuric acid injected through the closed rubber septum. ~4CO2 was collected in the center wells by the addition of 2 ml 2-ethoxyethanol/ethanolamine(1:1 vol/vol). Flasks were shaken for 30 rain at 110 rpm and center wells were removed and placed into 10 ml Aquasol. Radioactivity was counted as described above and quenching was determined using the channels ratio method. Sediment concentrations of glutamate and acetate concentrations were not determined. Therefore, the rates of J4CO2 formation are expressed as relative mineralization rates in contrast to the phenanthrene mineralization rates.
Chemicals [9-t4C]phenanthrene, specific activity 6.12 mCi/mM, was obtained from Amersham Co., Arlington Heights, IL. Radiolabeled phenanthrene was purified prior to use by Florisil (Fisher) column chromatography to a purity greater than 98%, as determined by TLC analysis. L-[U-~4C]glutamate, specific activity 1.4 mCi/mg, was obtained from Research Products International Corp., Mount Pleasant, IL. All solvents were distilled in glass (Burdick and Jackson, Muskegon, MI).
Statistical Analyses Linear regression analysis and analysis of variance were carded out on an IBM-PC/XT microcomputer using the STATPACK statistical package.
Results T w o t r a n s e c t e x p e r i m e n t s were c o n d u c t e d , o n e o n A p r i l 3 a n d t h e o t h e r o n A p r i l 28, 1984. O n t h e first date, p h e n a n t h r e n e m i n e r a l i z a t i o n a n d n o t glut a m a t e m i n e r a l i z a t i o n w a s m e a s u r e d ; h o w e v e r , o n A p r i l 28 b o t h a c t i v i t i e s were m e a s u r e d . T h e A p r i l 3 d a t a is n o t g i v e n b e c a u s e the r e l a t i v e site rates a n d s a l i n i t y effects were the s a m e i n b o t h e x p e r i m e n t s . T h e a v e r a g e site t e m p e r a t u r e o n A p r i l 3 was 6.4~ a n d t h e p h e n a n t h r e n e rates were a p p r o x i m a t e l y 7 5 % l o w e r t h a n the A p r i l 28 rates p r e s e n t e d below.
Transect Site Characteristics. P h y s i c a l a n d c h e m i c a l c h a r a c t e r i s t i c s o f the N e p o n s e t R i v e r t r a n s e c t sites are g i v e n i n T a b l e 1 for the t r a n s e c t e x p e r i m e n t t h a t was c o n d u c t e d o n A p r i l 28,
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Table 1. Physical and chemical characteristics o f the sediment sites along a transect in the Neponset River, Boston Harbor, Massachusetts
Site Salinity no, (%o) 1 2 3
0 15.0 30.0
Temperature (~ 10.2 10.3 10.3
pH
TOC (mg/ g)~
Phenanthrene (ng/g)a
7.9 8.1 8.1
23 19 25
775 (31) b 953 (80) 439 (30)
a Per g dry sediment ~'(Standard deviation) of three replicates
1984. T h e sites were chosen to give 0, 15, and 30%0 salinities near the s e d i m e n t / water interface for sites 1, 2, and 3, respectively. All three sites were similar in observed physical and chemical characteristics with the exception o f salinity. Concentrations o f p h e n a n t h r e n e were similar at all three sites, ranging from 439 to 953 ng/g dry sediment. T O C , pH, and t e m p e r a t u r e were also similar at the three sites, since all the sites were sampled within 2 hours.
Phenanthrene Mineralization Rates o f p h e n a n t h r e n e mineralization increased with increasing salinity in both natural and artificial seawater-amended sediment slurries (Fig. 2). T h e relationship as analyzed by linear regression analysis was significant for b o t h natural seawater and M S S - a m e n d e d slurries (P < 0.001). T h e calculated statistics for phenanthrene mineralization rate vs. a m b i e n t salinity were as follows: regression coefficient = 0.940, slope = 0.523, and y-intercept = 2.267 for natural water slurries, and regression coefficient = 0.950, slope = 0.402, and y-intercept = 1.967 for artificial water slurries. Rates for site 1 sediments (ambient salinity, 0%0) were not significantly affected by mixing slurries with 15 and 30%0 water salinities, either natural seawater or MSS. In contrast, p h e n a n t h r e n e rates at site 2, (ambient salinity, 157~) were significantly inhibited by b o t h 0%0 seawater and 07~ MSS, but not by waters at a higher salinity (30700) than the a m b i e n t site salinity. Similarly, p h e n a n t h r e n e mineralization rates at site 3, (ambient salinity, 307~), were significantly inhibited by 07~ seawater and MSS at 0 and 15%0 salinity. This is in contrast to p h e n a n t h r e n e mineralization at the freshwater site, site 1, which was tolerant to a wide range o f salinities.
Glutamate Mineralization G l u t a m a t e mineralization, as an estimate o f potential heterotrophi c activity in sediments, was affected by salinity in a m a n n e r similar to p h e n a n t h r e n e mineralization (Fig. 3). G l u t a m a t e mineralization rates were higher at sites 2 and 3 as c o m p a r e d to site 1; however, there was not a significant relationship to
Salinity and Phenanthrene Biodegradation
141 Z I.d
I.Z uJ
18 NATURAL SEAWATER 16 Z
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~o
ARTIFICIAL SEAWATER
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_o 9~
J4
_1 Z
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'"
6
IZ w
O
15 30
EXPERIMENTAL
0
15 30
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15 30
Fig. 2. Rates of phenanthrene mineralization by Neponset River sediment slurries from three sites along a salinity transect. Slurries were mixed with natural seawater or artificial seawater of varying salinity. Error bars are _+ standard deviation of three replicate sediment grabs.
SALINITY (%.)
ambient salinity as with p h e n a n t h r e n e mineralization. In part, this was due to the higher variability observed with glutamate rates as c o m p a r e d to p h e n a n threne rates. Varying salinity did not affect glutamate mineralization o f site 1 sediments. Only 0%0 MSS significantly inhibited glutamate activity in site 2 sediments (ambient salinity, 15%0). Site 3 sediments (ambient salinity, 307~) were significantly inhibited by both 0%0 seawater and 0%0 MSS. Therefore, as with phenanthrene mineralization, the heterotrophic microbial activity in the brackish sediment sites was inhibited by 0%0 water.
Discussion Salinity changes can strongly influence rates o f P A H degradation in m a r i n e sediments [22]. In a previous field study, I reported that naphthalene and p h e n a n t h r e n e b i o t r a n s f o r m a t i o n rates were significantly stimulated by increasing salinity during a 15-month period in estuarine sediments [36]. T h e results presented here confirm the influence o f salinity on p h e n a n t h r e n e degradation in brackish sediments and suggest that the p r e d o m i n a n t p h e n a n t h r e n e degraders in brackish water are obligate marine microorganisms. Experiments employing an artificial seawater (minimal salts solution, MSS) revealed that inhibition o f activities at low salinities was due p r e d o m i n a t e l y to the s o d i u m
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M . P . Shiaris
NATURAL SEAWATER m
ARTIFICIAL
SEAWATER
LU Iz Lt.I
Iz
Z
o_.... 12
~
~ ~,,9
N~ --I Z ,~ tt,J
I0
IZ u.I
o,,,
Looo
Fig. 3. Rates o f relative glutamate mineralization by Neponset River sediment slurries from three sites along a salinity transect. Slurries were m i x e d with natural seawater or artificial seawater o f varying salinity. Error bars are ___ standard deviation o f three replicate sediment grabs.
~m 4
2
5 0
15 ~0 EXPERIMENTAL
0
15 30 SALINITY
0
15 30
(%0)
chloride content of the seawater alone and not to differences in the chemistry of other seawater constituents. Though general microbial activity displayed similar trends, the effect of salinity was not as marked as compared to the salinity effect on phenanthrene mineralization. Phenanthrene-degrading bacteria represent a small subset of the total bacterial community in sediments [37], and therefore they probably contribute little to the overall heterotrophic activity in the sediments. The natural water slurries demonstrated consistently higher or equal phenanthrene-mineralizing activity as compared to the artificial water slurries. A similar trend was found with glutamate mineralization with two exceptions. Higher activity in natural water slurries indicates that the organic and perhaps inorganic constituents of seawater stimulate the heterotrophic activity of the extant microbial communities. The availability of trace vitamins, elements, and organics may stimulate microbial growth and perhaps increase the bioavailability of phenanthrene by increasing its solubility through surfactant action. In addition, small deviations in the artificial water salinity and pH from in situ values may have had an inhibitory effect on the activities. However, these constituents, absent in artificial water, were only a minor component of the observed slurry effects as the effect of artificial water-amended slurries on phenanthrene mineralization generally covaried with the effect of natural water amendments. PAH degraders have been isolated repeatedly from marine sediments. West et al. [44] described several phenanthrene-degrading strains of Vibrio, a genus in which most species require 2-3% NaC1 for optimum growth [24]. A similar phenanthrene degrader was described by Kiyohara and Nagao [23]. Vibrios, however, are more affiliated with the water column than sediments [10], and their role in PAH degradation has not been examined. Though many other
Salinity and Phenanthrene Biodegradation
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PAH degraders have been isolated from marine sediments [12, 15, 37, 39], their status as obligate marine bacteria has not been examined. The salinity regime of individual sediment sites appears to influence the effect of salinity changes on the P A H mineralizing activity. Kerr and Capone [22] reported that P A H degradation in a low salinity sediment from the upper reaches of the Hudson River was inhibited by higher salinity water. This was in contrast to a low salinity downstream site which was more tolerant to higher salinity water as observed in the Neponset River. Sediments in the relatively small Neponset River are subject to constant daily and seasonal fluctuations in salinity; particularly at the freshwater site. At the time of sampling, the river displayed a classic salt wedge profile with salinity increasing vertically with depth at sites 2 and 3. As the wedge moves back and forth with the tides and sporadic freshwater flow, the sediments are subjected to varying salinities. In general, site 1 is typically freshwater but experiences frequent episodes of saltwater intrusion, particularly during periods of low rainfall. Conversely, site 3 usually has a salinity of > 25%o except for infrequent periods of unusually heavy rainfall. It is doubtful that the salinity at site 3 ever drops to 0 ~ . Site 2 can be under extreme daily fluctuation of salinity ranging from 10 to 3 0 ~ which may explain why phenanthrene mineralization was optimal at both 15 and 30ff~ seawater. Apparently, the degraders at low salinities are well adapted to a fluctuating saline environment. Their salinity tolerance could have two underlying mechanisms: Either the individual degraders are tolerant to a wide range of salt concentrations, or the microbial consortium of phenanthrene degraders comprises a mixture of individual types, each with different salt tolerances. The results suggest the first hypothesis is more plausible, that individual degraders are euryhaline, because incubations were fairly short and lag times were not observed at salinitics different from ambient. In contrast to the relationship between phenanthrene degradation and salinity in the Neponset River, Readman et al. [34] did not find a significant correlation in the Tamar Estuary, England; however, the rates were examined only in the water column where the numbers of P A H degraders are likely to be minimal [37]. A significant salinity effect has been reported for the degradation of other aromatics and xenobiotic compounds. For these compounds, degradation rates decreased with increasing salinity [8, 40, 43]. Because the degrading populations are advected from upstream water, a direct comparison with presumably resident sediment populations cannot be made. Similarly, Bartholomew and Pfaender [3] found higher uptake rates of nitrilotriacetic acid, cresol, chlorobenzene, and trichlorobenzene in the freshwater sites of a natural salinity gradient. The reports of degradation potential as a function of salinity are conflicting, but degradation potential, to a large degree, can be attributed to the preexposure history of marine and estuarine communities to compounds [32, 41]. For example, adaptation to high concentrations of PAH results in significantly increased P A H degrading activity [5, 35]. In the case of the Neponset River, however, phenanthrene concentrations varied only about twofold from site to site and exposure to phenanthrene alone does not explain the linear increase of mineralization rates with increasing salinity. Phenanthrene mineralization rates at other Boston Harbor sites are directly correlated to ambient P A H
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M.P. Shiaris
concentrations [36]. Kerr and C a p o n e [22] described a similar relationship between P A H mineralization a n d a m b i e n t concentration, even at sediments o f disparate salinity regimes. It is not clear w h y salinity correlated with phenanthrene degradation in the N e p o n s e t River. Possible causes include marine degraders' intrinsically higher rates o f P A H mineralization or inhibition o f activity by potential toxic c o m p o u n d s in higher concentrations at the freshwater site. Salt concentration can also affect the interaction o f P A H with other sedi m e n t constituents [45], which in turn m i g h t affect P A H bioavailability to the degraders [42]. The effects o f particulates a n d organic c o m p o u n d s on the rates o f P A H degradation is an area that is poorly u n d e r s t o o d [30]. The existence o f obligate m a r i n e P A H degraders underscores the diversity o f bacteria in the e n v i r o n m e n t capable o f P A H degradation. Yet the elegant physiological and genetic studies that have yielded considerable i n f o r m a t i o n on the nature o f bacterial P A H degradation have focused primarily on a few strains o f Pseudomonas a n d a Beijerinckia isolated from p e t r o l e u m - c o n t a m i nated soils [ 14]. Future efforts in this l a b o r a t o r y will be to isolate a n d c o m p a r e marine isolates to the well-described soil degraders. These bacteria m a y also provide new genetic material for potential use in biological r e m e d i a t i o n o f contaminated environments.
Acknowledgments. This research was funded by Environmental Protection AgencyGrants R-81181801-0 and R-810119-01-0, and a Biomedical Research Support Grant from the University of Massachusetts at Boston. I thank D. Jambard-Sweet and C. Emmett for technical assistance and D. G. Capone and C. O'Rork for critique of the manuscript.
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11. DeLaune RD, Patrick WH, Casselman ME (1981) Effect of sediment pH and redox conditions on degradation of benzo(a)pyrene. Mar PoUut Bull 12:251-253 12. Garcia-Valdes E, Cosar E, Rotger R, Lalucat J, Ursing J (1988) New naphthalene-degrading marine Pseudomonas strains. Appl Environ Microbiol 54:2478-2485 13. Gardner WS, Lee RF, Tenore KR, Smith LW (1979) Degradation of selected polycyclie aromatic hydrocarbons in coastal sediments: Importance of microbes and polychaete worms. Water Air Soil Pollut 11:339-347 14. Gibson DT (1968) Microbial degradation of aromatic compounds. Science 161:1093-1097 15. Guerin WF, Jones GE (1988) Mineralization of phenanthrene by a Mycobacterium sp. Appl Environ Microbiol 54:937-944 16. Hambrick GA, DeLaune RD, Patrick WH (1980) Effect of estuarine sediment pH and oxidation-reduction potential on microbial hydrocarbon degradation. Appl Environ Microbiol 40:365-369 17. Heitkamp MA, Cemiglia CE (1987) Effects of chemical structure and exposure on the microbial degradation of polycyclic aromatic hydrocarbons in freshwater and estuarine ecosystems. Environ Toxicol Chem 6:535-546 18. Heitkamp MA, Freeman JP, Cerniglia CE (1987) Naphthalene biodegradation in environmental microcosms: Estimates of degradation rates and characterization of metabolites. Appl Environ Microbiol 53:129-136 19. Herbes SE, Schwall LR (1978) Microbial transformation ofpolycyclic aromatic hydrocarbons in pristine and petroleum-contaminated sediments. Appl Environ Microbiol 35:306-316 20. Hites RA, Biemann K (1972) Water pollution: Organic compounds in the Charles River, Boston. Science 178:158-160 21. Hites RA, LaFlammme RE, Farrington JW (1977) Polycyclic aromatic hydrocarbons in recent sediments: The historical record. Science 198:829-831 22. Kerr RP, Capone DG (in press) The effect of salinity on the microbial mineralization of two polycyclic aromatic hydrocarbons in estuarine sediments. Mar Environ Res 23. Kiyohara H, Nagao K (1978) The catabolism of phenanthrene and naphthalene by bacteria. J Gen Microbiol 105:69-75 24. Ka'ieg NR~, Holt JG (eds) (1984) Bergey's manual of systematic bacteriology, vol I. Williams and Wilkins, Baltimore 25. Lee RF, Ryan C (1976) Biodegradation of petroleum hydrocarbons by marine microbes. In: Sharpley JM, Kaplan AM (eds) Proc 3rd Int Biodegradation Symposium. Applied Science Publishers, London, pp 119-125 26. Leversee GJ, Giesy JP, Landrum PF, Gerould S, Bowling JW, Fanwin TE, Haddock JD, Bartell SM (1982) Kinetics and biotransformation of benzo(a)pyrene in Chironomus riparius. Arch Environ Contain Toxicol I 1:25-32 27. Levinton JS (1982) Marine ecology. Prentice-Hall, Inc, Englewood Cliffs, New Jersey 28. May WE (1980) The solubility behavior of polycyclic aromatic hydrocarbons in aqueous systems. In: Pctrakis L, Weiss FT (eds) Petroleum in the marine environment, American Chemical Society, Washington DC 29. McElroy AE (1985) In vivo metabolism of benz(a)anthracene by the polychaete Nereis virens. Mar Environ Res 17:133-136 30. MeElroy AE, Farrington JW, Teal JM (1989) Metabolism of polynuclear aromatic hydrocarbons in the aquatic environment. In: Varanasi U (ed) Metabolism of polynuclear aromatic hydrocarbons in the aquatic environment. CRC Press, Boca Raton, pp 1-39 31. Means JC, Hasett JJ, Wood SG, Banwart WL (1980) Sorption properties of polynuclear aromatic hydrocarbons by sediments and soils. Environ Sei Technol 14:1524-1528 32. Pfaender FK, Shimp RJ, Larson RJ (1985) Adaptation of estuarine ecosystems to the biodegradation of nitrilotriacetic acid: Effects of preexposure. Environ Toxicol Chem 4:587-593 33. Pipe RK, Moore MN (1986) An ultrastructural study on the effects of phenanthrene on lysosomal membranes and distribution of the lysosomal enzyme fl-glucuronidase in digestive ceils of the periwinkle Littorina littorea. Aquat Toxicol 8:65-76 34. Readman JW, Mantoura RFC, Rhead MM, Brown L (1982) Aquatic distribution and hetero-
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M.P. Shiaris trophic degradation of polycyclic aromatic hydrocarbons (PAH) in the Tamar Estuary. Estuar Coast Shelf Sci 14:369-389 Sayler GS, Perkins RE, Sherrill TW, Perkins BK, Reid MC, Shields MS, Kong HL, Davis JW (1983) Microcosm and experimental pond evaluation of microbial community response to synthetic oil contamination in freshwater sediments. Appl Environ Microbiol 46:211-219 Shiaris MP (1989) Seasonal biotransformation of naphthalene, phenanthrene, and benzo[a]pyrene in surficial estuarine sediments. Appl Environ Microbiol 55:1391-1399 Shiaris MP, Cooney JJ (1983) Replica plating method for estimating phenanthrene-utilizing and phenanthrene-cometabolizing microorganisms. Appl Environ Microbiol 45:706-710 Shiaris MP, Jambard-Sweet D (1986) Polycyclic aromatic hydrocarbons in surficial sediments of Boston Harbour, Massachusetts, USA. Mar Pollut Bull 17:469--472 Sisler FD, Zobell CE (1947) Microbial utilization of carcinogenic hydrocarbons. Science 106: 521-522 Spain JC, Pritchard PH, Bourquin AW (1980) Effects of adaptation on biodegradation rates in sediment/water cores from estuarine and freshwater environments. Appl Environ Microbiol 40:726-734 Spain JC, Van Veld PA (1983) Adaptation of natural microbial communities to degradation of xenobiotic compounds: Effects of concentration, exposure time, inoculum, and chemical structure. Appl Environ Microbiol 45:428-435 Varanasi U, Reichert WL, Stein JE, Brown DW, Sanborn HR (1985) Bioavailability and biotransformation of aromatic hydrocarbons in benthic organisms exposed to sediment from an urban estuary. Environ Sci Technol 19:836-841 Ward DM, Brock TD (1978) Hydrocarbon biodegradation in hypersaline environments. Appl Environ Microbiol 35:353-359 West PA, Okpokwasili GC, Brayton PR, Grimes DJ, Colwell RR (1984) Numerical taxonomy ofphenanthrene-degrading bacteria isolated from the Chesapeake Bay. Appl Environ Microbiol 48:988-993 Wijayaratne RD, Means JC (1984) Sorption of polycyclic aromatic hydrocarbons by natural estuarine colloids. Mar Environ Res 11:77-89 Wolfe DE (ed) (1977) Fates and effects of petroleum hydrocarbons in marine organisms and ecosystems. Pergamon Press, New York Youngblood MW, Blumer M (1975) Polycyclic aromatic hydrocarbons in the environment: Homologous series in soils and recent marine sediments. Geochim Cosmochim Acla 39:13031314
MICROBIAL ECOLOGY | Springer-Verlag New York Inc. 1989
Phenanthrene Mineralization Along a Natural Salinity Gradient in an Urban Estuary, Boston Harbor, Massachusetts Michael P. Shiaris University of Massachusetts at Boston, Biology Department, Boston, Massachusetts 02125, USA
Abstract. The effect of varying salinity on phenanthrene and glutamate mineralization was examined in sediments along a natural salinity gradient in an urban tidal river. Mineralization was measured by trapping ~4CO2 from sediment slurries dosed with trace levels of [14C]phenanthrene or [14C]glutamate. Sediments from three sites representing three salinity regimes (0, 15, and 30%0) were mixed with filtered column water from each site. Ambient phenanthrene concentrations were also determined to calculate phenanthrene mineralization rates. Rates of phenanthrene mineralization related significantly to increasing salinity along the transect as determined by linear regression analysis. Rates ranged from 1 ng/hour/g dry sediment at the freshwater site to > 16 ng/hour/g dry sediment at the 30%00salinity site. Glutamate mineralization also increased from the freshwater tO the marine site; however, the relationship to salinity was not statistically significant. To examine the effect of salinity on mineralizing activities, individual sediments were mixed with filtered water of the other two sites. Slurries were also made with artificial seawater composed of 0, 15, or 30 g NaCI/ liter to substitute for overlying water. Rates ofphenanthrene mineralization in the 0%0 ambient salinity sediments were not affected by higher salinity waters. Activities in the 15 and 30%0 ambient salinity sediments, however, were significantly inhibited by incubation with 0%0 salinity water. The inhibition, in large part, appears to be due to the decreased NaC1 concentration of the water phase. Glutamate mineralization was affected in a similar manner, but not as dramatically as phenanthrene mineralization. The results suggest that phenanthrene degraders in low salinity estuarine sediments subject to salt water intrusion are tolerant to a wide range of salinities but phenanthrene degradation in brackish waters is mainly a function of obligate marine microorganisms.
Introduction
Polycyclic aromatic hydrocarbons (PAHs), a class of compounds composed of two or more fused benzene rings, are of concern to marine biota and h u m a n public health as toxic, mutagenic, and carcinogenic agents [2, 46]. Phenan-
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threne, a three-ring PAH, is not a mutagen but exhibits toxicity to marine organisms [7, 33]. Compared with higher molecular weight PAHs, phenanthrene is readily biodegraded and serves as a useful model for studying the fate of PAHs in the environment. Individual P A H compounds have low water solubilities [28] which decrease with increasing molecular weight. Consequently, they display a high affinity for organic matter and particles in aquatic ecosystems [31]. With particle settlement, PAHs, including phenanthrene, often reach high levels in aquatic sediments neighboring sources of PAH-generating activities. Although sediments are exposed to naturally derived PAH, it is anthropogenic sources that contribute the predominant share of PAH to urban waters. For example, the depositional history of industrial and energy-generated PAH inputs can be readily chronicled in undisturbed sediments adjacent to urban areas [21]. Urban estuaries, such as Boston Harbor, attain P A H concentrations exceeding 100 mg/g sediment [20, 38, 47]. Once trapped in aquatic sediments, PAHs persist for long periods, and, depending on environmental conditions and the structure of the compound, they may last indeterminately. Chemical oxidation and photooxidation may be important routes of P A H loss in the water column, but biological transformation is probably the prevailing mechanism of P A H loss in sediments. The biological participants include many eukaryotic organisms native to marine sediments [26, 29, 42] as well as a diverse group of bacteria isolated from marine environments [12, 15, 23, 37, 39, 44]. Many of the bacteria, in contrast to the eukaryotes, are capable of utilizing PAHs as sole carbon and energy sources [9]. The importance of microorganisms in the degradation and fate of PAHs in marine sediments has been well documented [4, 17, 18, 25]. The fate of P A H in sediments is under the control of complex biological and environmental factors [1]. For example, oxygen is a master environmental variable that controls degradation rates [11, 16]. Both eukaryotic and prokaryotic degradation pathways utilize bimolecular oxygen in the initial enzymatic attack on the P A H ring [9], yet oxygen is often limiting in sediments. The influx of oxygen into sediments can be augmented by burrowing animals which in turn stimulate microbial degradation of P A H [6, 13]. Furthermore, the degree of mixing of oxygen into anaerobic sediments is influenced by the numbers, types, and seasonal succession of benthic animals in the sediments [27]. In addition to oxygen, P A H biodegradation is affected by nutrients, temperature [1], microbial adaptation to PAHs [35], and the composition of the PAH mixture [5]. In marine environments, salinity also appears to have a significant effect on microbial degradation of PAHs. Kerr and Capone [22] reported that naphthalene and anthracene degraders in the Hudson River estuary were highly adapted to their salinity regime. I observed that degradation of naphthalene and phenanthrene in three urban estuarine sediments was significantly related to seasonal salinity changes over the course of a 15-month period [36]. The same trend was observed for benzo(a)pyrene degradation for one of the sites. These results suggested that P A H degradation was carried out predominantly by obligate marine microorganism. The objective of the present study was to determine if salinity affects phenanthrene mineralization along a natural salinity gradient and whether mineralization is dependent on the presence of sodium chloride.
Salinity and Phenanthrene Biodegradation
137
GREATER BOSTON AREA
BOSTON HARBOR UMoSs
0------9
50%= SITE
m. ~
0
0,5
!
I
1 IKm
15%. SITE 0%= ;ITE
Fig. I. Map of the Neponset River, a tributary of Boston Harbor, Massachusetts. The three sampling sites are located within the tidal extension of the river. Methods
Sampling Three sites along a transect in the tidal Neponset River were chosen to represent a salinity gradient ranging from 0 to 30%0 (Fig. 1). The Neponset River is a tributary of Boston Harbor, Massachusetts. Boston Harbor is a glacially carved, tidally dominated estuary, typical of many bays in New England. The Harbor and its 31 associated islands cover an area of 114 square kilometers with 190 km of shoreline. All reaches of the Harbor and its tributary streams are polluted (Federal Water Pollution Control Federation, Northeast Water Quality Management Center, 1967, Department of the Interior, Washington DC), and at least one-third of the harbor is grossly polluted by municipal and industrial sewage with deposits of decayed organic matter and oil residues covering much of the bottom (M. G. Fitzgerald, 1980, Woods Hole Oceanographic Institute, WHOI-80-38, Woods Hole, MA). Freshwater influx to the Harbor is low and the entire Harbor is flushed in 42 days, predominantly by diurnal tides. Aside from two main shipping channels, the mean depth of the harbor is 3 m or less at mean low water. All three sites were composed of muds consisting of fine silt-clays. Sediment samples were taken in triplicate at each site with an Eckman dredge (Wildco, Saginaw, Michigan). The top 2 - 3 - m m sediment of each grab was subsampled with a sterile wooden spatula,
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placed in glass bottles, kept at ambient water temperature in chest coolers, and returned to the laboratory for immediate processing. Salinity and conductivity (Beckman RS5-3 Salinometer), sediment temperature, and pH were recorded at each site and sampling date. Total organic carbon by sediment acidification was determined on a C H N analyzer by the Darling Marine Laboratory, Walpole, Maine. Ambient phenanthrene concentrations were determined by Soxhlet extraction of sediments with methanol and methanol/benzene, column chromatography fractionation, and high performance liquid chromatography/diode array detection analysis, as described previously [38].
Phenanthrene Mineralization Rates Techniques for determining PAH mineralization rates were based on the protocol of Herbes and Schwall [19]. Surficial sediment samples, one from each sediment grab, were diluted 1:10 (wt/wt) in estuarine water or autoclaved m i n i m u m salts solution (MSS) and homogenized with a magnetic stirrer. The salinity of the estuarine water or MSS was either 0, 15, or 30%0. MSS was composed of 0.8 g K2HPO4, 0.2 g KH2PO4, 0.05 g CaSO4-2H20, 0.5 g MgSO4-7H20, 1.0 g (NH4)2SO4,0.01 g FeSO4"2H20, and 1 liter distilled water (pH 8.0) and amended with either 0, 15, or 30 g NaC1/ liter final concentration. The estuarine water was obtained from the water column overlying th~ sampling site and filtered through a 0.2-#m Nuclepore membrane prior to mixing. Aliquots of the sediment slurry (10 ml) were transferred via sterile, wide-mouth, serological pipets to sterile 25ml Erlenmeyer flasks, one for each sediment grab, therefore three per site. The flasks were fitted with Teflon-lined serum caps and glass center wells. Aliquots (0.1 ml) ofradiolabeled PAH dissolved in acetone were added to each incubation flask at a final concentration of 20 ng/g wet slurry, a concentration below the ambient PAH concentrations. Flasks were incubated at in situ temperatures in the dark without shaking for 60 hours. During the incubation, sediments settled to the bottom of the flask at an approximate thickness of 1 ram. No discernable change in the light color of the sediments occurred during incubation indicating that the sediments remained oxidized. Mineralization rates were linear with time (data not shown), which also indicated that oxygen was not depleted. The incubation time resulted in less than 10% transformation of the parent radiolabeled phenanthrene. All experiments included frozen controls, which were identical to occasional controls killed by autoclaving or formalin (2% final concentration) addition. Incubations were terminated by injecting 0.3 ml 10% trichloroacetic acid through the Teflon-lined serum cap. Radiolabeled CO2 was trapped in the glass center wells by the addition of 0.3 ml 2 N K O H to the wells. K O H was used rather than an organic solvent CO2-trapping agent to minimize contamination with radiolabeled phenanthrene. Flasks were shaken for 30 m i n at 1 I0 rpm and the K O H was transferred into vials containing 10 ml Aquasol (New England Nuclear, Boston, MA) for liquid scintillation counting. Counts were performed on a Packard Model 3330 Tricarb liquid scintillation counter. Quenching was determined by the internal standards method employing [t4C]glutamate for the Aquasol cocktails. Preliminary experiments were conducted to ensure that the center well KOH was free of [~4C]phenanthrene. The KOH was extracted with benzene and examined by thin-layer chromatography (TLC) and liquid scintillation counting. Phenanthrene could not be detected on TLC and scintillation counts were less than 0.1% of the [14C]phenanthrene counts added to the flasks. Mineralization rates were calculated with the assumption that added radioactive PAH was completely mixed and equilibrated with ambient sediment PAH. The mineralization rate, R, was calculated as follows: R = r (Ca + C~)/Cr where, r = rate of [14C]phenanthrene mineralization corrected for sterile controls Ca = ambient concentration of PAH in sediment C, = concentration of added ~4C-phenanthrene
Salinity and Phenanthrene Biodegradation
139
Because [~4C]phenanthrene was added in trace amounts compared to known ambient phenanthrene concentrations, R represents an accurate estimate of mineralization rates in the sediment slurries.
Glutamate Mineralization Incubationflasks were prepared as for phenanthrene mineralizationassays except that plastic center wells and rubber septa without Teflon linings were used. Radiolabeled glutamate was added at a final concentration of 130 pg/ml of slurry. Incubations were carded out in the dark at in situ temperatures for 1 hour. Replicate controls consisted of sediment slurries killed with 0.5 ml m-cresol. Incubations were terminated by the addition of 0.3 ml 2 M sulfuric acid injected through the closed rubber septum. ~4CO2 was collected in the center wells by the addition of 2 ml 2-ethoxyethanol/ethanolamine(1:1 vol/vol). Flasks were shaken for 30 rain at 110 rpm and center wells were removed and placed into 10 ml Aquasol. Radioactivity was counted as described above and quenching was determined using the channels ratio method. Sediment concentrations of glutamate and acetate concentrations were not determined. Therefore, the rates of J4CO2 formation are expressed as relative mineralization rates in contrast to the phenanthrene mineralization rates.
Chemicals [9-t4C]phenanthrene, specific activity 6.12 mCi/mM, was obtained from Amersham Co., Arlington Heights, IL. Radiolabeled phenanthrene was purified prior to use by Florisil (Fisher) column chromatography to a purity greater than 98%, as determined by TLC analysis. L-[U-~4C]glutamate, specific activity 1.4 mCi/mg, was obtained from Research Products International Corp., Mount Pleasant, IL. All solvents were distilled in glass (Burdick and Jackson, Muskegon, MI).
Statistical Analyses Linear regression analysis and analysis of variance were carded out on an IBM-PC/XT microcomputer using the STATPACK statistical package.
Results T w o t r a n s e c t e x p e r i m e n t s were c o n d u c t e d , o n e o n A p r i l 3 a n d t h e o t h e r o n A p r i l 28, 1984. O n t h e first date, p h e n a n t h r e n e m i n e r a l i z a t i o n a n d n o t glut a m a t e m i n e r a l i z a t i o n w a s m e a s u r e d ; h o w e v e r , o n A p r i l 28 b o t h a c t i v i t i e s were m e a s u r e d . T h e A p r i l 3 d a t a is n o t g i v e n b e c a u s e the r e l a t i v e site rates a n d s a l i n i t y effects were the s a m e i n b o t h e x p e r i m e n t s . T h e a v e r a g e site t e m p e r a t u r e o n A p r i l 3 was 6.4~ a n d t h e p h e n a n t h r e n e rates were a p p r o x i m a t e l y 7 5 % l o w e r t h a n the A p r i l 28 rates p r e s e n t e d below.
Transect Site Characteristics. P h y s i c a l a n d c h e m i c a l c h a r a c t e r i s t i c s o f the N e p o n s e t R i v e r t r a n s e c t sites are g i v e n i n T a b l e 1 for the t r a n s e c t e x p e r i m e n t t h a t was c o n d u c t e d o n A p r i l 28,
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M . P . Shiaris
Table 1. Physical and chemical characteristics o f the sediment sites along a transect in the Neponset River, Boston Harbor, Massachusetts
Site Salinity no, (%o) 1 2 3
0 15.0 30.0
Temperature (~ 10.2 10.3 10.3
pH
TOC (mg/ g)~
Phenanthrene (ng/g)a
7.9 8.1 8.1
23 19 25
775 (31) b 953 (80) 439 (30)
a Per g dry sediment ~'(Standard deviation) of three replicates
1984. T h e sites were chosen to give 0, 15, and 30%0 salinities near the s e d i m e n t / water interface for sites 1, 2, and 3, respectively. All three sites were similar in observed physical and chemical characteristics with the exception o f salinity. Concentrations o f p h e n a n t h r e n e were similar at all three sites, ranging from 439 to 953 ng/g dry sediment. T O C , pH, and t e m p e r a t u r e were also similar at the three sites, since all the sites were sampled within 2 hours.
Phenanthrene Mineralization Rates o f p h e n a n t h r e n e mineralization increased with increasing salinity in both natural and artificial seawater-amended sediment slurries (Fig. 2). T h e relationship as analyzed by linear regression analysis was significant for b o t h natural seawater and M S S - a m e n d e d slurries (P < 0.001). T h e calculated statistics for phenanthrene mineralization rate vs. a m b i e n t salinity were as follows: regression coefficient = 0.940, slope = 0.523, and y-intercept = 2.267 for natural water slurries, and regression coefficient = 0.950, slope = 0.402, and y-intercept = 1.967 for artificial water slurries. Rates for site 1 sediments (ambient salinity, 0%0) were not significantly affected by mixing slurries with 15 and 30%0 water salinities, either natural seawater or MSS. In contrast, p h e n a n t h r e n e rates at site 2, (ambient salinity, 157~) were significantly inhibited by b o t h 0%0 seawater and 07~ MSS, but not by waters at a higher salinity (30700) than the a m b i e n t site salinity. Similarly, p h e n a n t h r e n e mineralization rates at site 3, (ambient salinity, 307~), were significantly inhibited by 07~ seawater and MSS at 0 and 15%0 salinity. This is in contrast to p h e n a n t h r e n e mineralization at the freshwater site, site 1, which was tolerant to a wide range o f salinities.
Glutamate Mineralization G l u t a m a t e mineralization, as an estimate o f potential heterotrophi c activity in sediments, was affected by salinity in a m a n n e r similar to p h e n a n t h r e n e mineralization (Fig. 3). G l u t a m a t e mineralization rates were higher at sites 2 and 3 as c o m p a r e d to site 1; however, there was not a significant relationship to
Salinity and Phenanthrene Biodegradation
141 Z I.d
I.Z uJ
18 NATURAL SEAWATER 16 Z
to') if)
a~
m
~o
ARTIFICIAL SEAWATER
1-
_o 9~
J4
_1 Z
U.l 8
'"
6
IZ w
O
15 30
EXPERIMENTAL
0
15 30
O
15 30
Fig. 2. Rates of phenanthrene mineralization by Neponset River sediment slurries from three sites along a salinity transect. Slurries were mixed with natural seawater or artificial seawater of varying salinity. Error bars are _+ standard deviation of three replicate sediment grabs.
SALINITY (%.)
ambient salinity as with p h e n a n t h r e n e mineralization. In part, this was due to the higher variability observed with glutamate rates as c o m p a r e d to p h e n a n threne rates. Varying salinity did not affect glutamate mineralization o f site 1 sediments. Only 0%0 MSS significantly inhibited glutamate activity in site 2 sediments (ambient salinity, 15%0). Site 3 sediments (ambient salinity, 307~) were significantly inhibited by both 0%0 seawater and 0%0 MSS. Therefore, as with phenanthrene mineralization, the heterotrophic microbial activity in the brackish sediment sites was inhibited by 0%0 water.
Discussion Salinity changes can strongly influence rates o f P A H degradation in m a r i n e sediments [22]. In a previous field study, I reported that naphthalene and p h e n a n t h r e n e b i o t r a n s f o r m a t i o n rates were significantly stimulated by increasing salinity during a 15-month period in estuarine sediments [36]. T h e results presented here confirm the influence o f salinity on p h e n a n t h r e n e degradation in brackish sediments and suggest that the p r e d o m i n a n t p h e n a n t h r e n e degraders in brackish water are obligate marine microorganisms. Experiments employing an artificial seawater (minimal salts solution, MSS) revealed that inhibition o f activities at low salinities was due p r e d o m i n a t e l y to the s o d i u m
142
M . P . Shiaris
NATURAL SEAWATER m
ARTIFICIAL
SEAWATER
LU Iz Lt.I
Iz
Z
o_.... 12
~
~ ~,,9
N~ --I Z ,~ tt,J
I0
IZ u.I
o,,,
Looo
Fig. 3. Rates o f relative glutamate mineralization by Neponset River sediment slurries from three sites along a salinity transect. Slurries were m i x e d with natural seawater or artificial seawater o f varying salinity. Error bars are ___ standard deviation o f three replicate sediment grabs.
~m 4
2
5 0
15 ~0 EXPERIMENTAL
0
15 30 SALINITY
0
15 30
(%0)
chloride content of the seawater alone and not to differences in the chemistry of other seawater constituents. Though general microbial activity displayed similar trends, the effect of salinity was not as marked as compared to the salinity effect on phenanthrene mineralization. Phenanthrene-degrading bacteria represent a small subset of the total bacterial community in sediments [37], and therefore they probably contribute little to the overall heterotrophic activity in the sediments. The natural water slurries demonstrated consistently higher or equal phenanthrene-mineralizing activity as compared to the artificial water slurries. A similar trend was found with glutamate mineralization with two exceptions. Higher activity in natural water slurries indicates that the organic and perhaps inorganic constituents of seawater stimulate the heterotrophic activity of the extant microbial communities. The availability of trace vitamins, elements, and organics may stimulate microbial growth and perhaps increase the bioavailability of phenanthrene by increasing its solubility through surfactant action. In addition, small deviations in the artificial water salinity and pH from in situ values may have had an inhibitory effect on the activities. However, these constituents, absent in artificial water, were only a minor component of the observed slurry effects as the effect of artificial water-amended slurries on phenanthrene mineralization generally covaried with the effect of natural water amendments. PAH degraders have been isolated repeatedly from marine sediments. West et al. [44] described several phenanthrene-degrading strains of Vibrio, a genus in which most species require 2-3% NaC1 for optimum growth [24]. A similar phenanthrene degrader was described by Kiyohara and Nagao [23]. Vibrios, however, are more affiliated with the water column than sediments [10], and their role in PAH degradation has not been examined. Though many other
Salinity and Phenanthrene Biodegradation
143
PAH degraders have been isolated from marine sediments [12, 15, 37, 39], their status as obligate marine bacteria has not been examined. The salinity regime of individual sediment sites appears to influence the effect of salinity changes on the P A H mineralizing activity. Kerr and Capone [22] reported that P A H degradation in a low salinity sediment from the upper reaches of the Hudson River was inhibited by higher salinity water. This was in contrast to a low salinity downstream site which was more tolerant to higher salinity water as observed in the Neponset River. Sediments in the relatively small Neponset River are subject to constant daily and seasonal fluctuations in salinity; particularly at the freshwater site. At the time of sampling, the river displayed a classic salt wedge profile with salinity increasing vertically with depth at sites 2 and 3. As the wedge moves back and forth with the tides and sporadic freshwater flow, the sediments are subjected to varying salinities. In general, site 1 is typically freshwater but experiences frequent episodes of saltwater intrusion, particularly during periods of low rainfall. Conversely, site 3 usually has a salinity of > 25%o except for infrequent periods of unusually heavy rainfall. It is doubtful that the salinity at site 3 ever drops to 0 ~ . Site 2 can be under extreme daily fluctuation of salinity ranging from 10 to 3 0 ~ which may explain why phenanthrene mineralization was optimal at both 15 and 30ff~ seawater. Apparently, the degraders at low salinities are well adapted to a fluctuating saline environment. Their salinity tolerance could have two underlying mechanisms: Either the individual degraders are tolerant to a wide range of salt concentrations, or the microbial consortium of phenanthrene degraders comprises a mixture of individual types, each with different salt tolerances. The results suggest the first hypothesis is more plausible, that individual degraders are euryhaline, because incubations were fairly short and lag times were not observed at salinitics different from ambient. In contrast to the relationship between phenanthrene degradation and salinity in the Neponset River, Readman et al. [34] did not find a significant correlation in the Tamar Estuary, England; however, the rates were examined only in the water column where the numbers of P A H degraders are likely to be minimal [37]. A significant salinity effect has been reported for the degradation of other aromatics and xenobiotic compounds. For these compounds, degradation rates decreased with increasing salinity [8, 40, 43]. Because the degrading populations are advected from upstream water, a direct comparison with presumably resident sediment populations cannot be made. Similarly, Bartholomew and Pfaender [3] found higher uptake rates of nitrilotriacetic acid, cresol, chlorobenzene, and trichlorobenzene in the freshwater sites of a natural salinity gradient. The reports of degradation potential as a function of salinity are conflicting, but degradation potential, to a large degree, can be attributed to the preexposure history of marine and estuarine communities to compounds [32, 41]. For example, adaptation to high concentrations of PAH results in significantly increased P A H degrading activity [5, 35]. In the case of the Neponset River, however, phenanthrene concentrations varied only about twofold from site to site and exposure to phenanthrene alone does not explain the linear increase of mineralization rates with increasing salinity. Phenanthrene mineralization rates at other Boston Harbor sites are directly correlated to ambient P A H
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concentrations [36]. Kerr and C a p o n e [22] described a similar relationship between P A H mineralization a n d a m b i e n t concentration, even at sediments o f disparate salinity regimes. It is not clear w h y salinity correlated with phenanthrene degradation in the N e p o n s e t River. Possible causes include marine degraders' intrinsically higher rates o f P A H mineralization or inhibition o f activity by potential toxic c o m p o u n d s in higher concentrations at the freshwater site. Salt concentration can also affect the interaction o f P A H with other sedi m e n t constituents [45], which in turn m i g h t affect P A H bioavailability to the degraders [42]. The effects o f particulates a n d organic c o m p o u n d s on the rates o f P A H degradation is an area that is poorly u n d e r s t o o d [30]. The existence o f obligate m a r i n e P A H degraders underscores the diversity o f bacteria in the e n v i r o n m e n t capable o f P A H degradation. Yet the elegant physiological and genetic studies that have yielded considerable i n f o r m a t i o n on the nature o f bacterial P A H degradation have focused primarily on a few strains o f Pseudomonas a n d a Beijerinckia isolated from p e t r o l e u m - c o n t a m i nated soils [ 14]. Future efforts in this l a b o r a t o r y will be to isolate a n d c o m p a r e marine isolates to the well-described soil degraders. These bacteria m a y also provide new genetic material for potential use in biological r e m e d i a t i o n o f contaminated environments.
Acknowledgments. This research was funded by Environmental Protection AgencyGrants R-81181801-0 and R-810119-01-0, and a Biomedical Research Support Grant from the University of Massachusetts at Boston. I thank D. Jambard-Sweet and C. Emmett for technical assistance and D. G. Capone and C. O'Rork for critique of the manuscript.
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